Archive for December, 2010

More on Human Experimentation

December 27, 2010

from
http://www.mit.edu/~dmaze/human_experimentation.html
Atomic Energy Commission Radiation Experiments
In December, 1993, Secretary of Energy Hazel O’Leary made a disturbing announcement: since the 1940’s, the U. S. Atomic Energy Commission had been sponsoring a series of tests on the effects of radiation on the human body. American citizens who had checked into hospitals for a variety of ailments had been secretly injected with varying amounts of plutonium and other radioactive materials without their knowledge. Most patients thought it was “just another injection,” but the secret studies left enough radioactive material in the patients’ bodies to readily induce cancer.
Unlike the Tuskegee study, the researchers here were careful to follow up on their patients, and in many cases were able to determine where the plutonium had spread to after the patients’ death several years later. However, its methods are arguably more ethically questionable than those used in the Tuskegee study. In the AEC tests, the patients’ trust in their doctors and hospitals had essentially been betrayed. While the patients did receive the treatment they needed for their initial ailment, they were being given something whose effects were wholly unknown but was expected to cause harm.
Other experiments were conducted with other agencies. “Human Experiment 133” tested the effects upon the pilot and crew of an aircraft flying through a radiation cloud for 25-40 minutes. The Department of Defense ran tests on people living downwind of atomic tests. Prisoners in Washington and Oregon were exposed to radiation with only minimal consent. The true purposes of these tests were kept hidden from the public and the test subjects for almost 50 years.
Related Sites
http://tis-nt.eh.doe.gov/ohre/roadmap/achre/: The fundamental government document about this radiation testing is the Final Report of the Advisory Committee on Human Radiation Experiments (the ACHRE Report). The Department of Energy set up this site for ACHRE. It includes a description of what ACHRE is and why it was established along with the entire ACHRE Report. This site also has a great deal of background material on the subject of radiation in general, along with a look at the history of radiation research since World War II. AHCRE also looks directly at the ethical challenges presented by radiation work, recounting the decisions of the researchers rather than coming to any conclusions on its own.
This site is well-organized and has a huge amount of information. The main DoE Human Radiation Experiments page contains links to most of this information, including a listing of all of the radiation experiments, oral histories of some of the participants, and a listing of declassified DoE documents about the radiation experiments.
http://www.brown.edu/Courses/Bio_Community_Health168C/achrecri.html: In Ethical Aerobics: ACHRE’s Flight From Responsibility, four doctors argue that, while the ACHRE Report was fairly complete, the rights of the experimental subjects were largely ignored. Early on in the paper, they point out that neither patients nor patient representatives were on ACHRE. They then examine the remainder of the Report, ultimately concluding that AHCRE’s presentation of the radiation testing was designed to absolve the U. S. government of responsibility for the experiments as much as possible.
http://www.brown.edu/Courses/Bio_Community_Health168C/consent.html: Michael Gipstein looks at the idea of informed consent in the ACHRE Report in his paper, The Historical Story of Informed Consent for Subjects of Radiation Experiments. He explores many specific examples of human experiments performed under AEC sanction, and is quick to point out the ethical lapses in each case.
http://www.yvwiiusdinvnohii.net/~nlthomas/political/sudradat.htm: Nancy Thomas maintains The People’s Paths, a site dedicated to Native American issues. She includes this article about a lawsuit against the U. S. government on behalf of Native Americans living near the Hanford, Washington nuclear reactor. The suit claims that residents of this area were systematically exposed to ionizing radiation and later tested without their knowledge.
http://www.webcom.com/~pinknoiz/coldwar/: This page has a couple of useful documents about radiation testing. An excerpt from the Intermediate Report of ACHRE details radiation experiments conducted by the CIA. Dr. David Egilman’s testimony about experiments conducted by the Department of Energy suggests that, just as in the Tuskegee experiment, treatment for radiation sickness was intentionally withheld from experimental subjects. Other documents here include a press release from the American Institute of Physics and some information on the Gulf War.
http://www.doe.gov/html/secretry/inside.html: Secretary of Energy Hazel O’Leary testified before the Senate Committee on Governmental Affairs in January, 1994. This excerpt from her testimony shows O’Leary doing her best to reassure Congress that, despite 40 years of announced experiments, the Department of Energy is doing its best to act ethically. O’Leary describes the drive to declassify documents about human radiation testing, and goes on to detail current testing, noting that “in no case is there any exposure of living human subjects to radioactive or chemical agents.” At the end of this excerpt, O’Leary announces the creation of ACHRE, a committee with a focus on “ethical and scientific standards.”
Other Links
http://www.breakpoint.org/scripts/70114.htm: This is a transcript from the radio show BreakPoint, described by its web page as “Your Daily Guide to Developing a Christian Worldview in a Post-Christian Culture”. Chuck Colson, the show’s narrator, argues on 14 January 1997 that new FDA rules are essentially a step back to Nazi experimentation. New rules allow a doctor to try an experimental treatment on an unconscious patient without his or her consent. BreakPoint states that the new rule is the FDA “caving in” to researchers’ complaints that it was too difficult to find patients willing to submit to experimental treatments and that it took too long to get an unconscious patient’s family’s approval to try a new treatment. Although its conclusions are somewhat extreme, BreakPoint’s argument against the new rules does stand on solid moral and religious ground.
Conclusions
During the 50 years since World War II, the U. S. government consistently ignored the ethical standards it set at the Nuremberg Trials for human experimentation to try to press through its perceived testing needs without alarming the American public. Whatever the gains, the ethical lapses far outshadowed the possible benefits from the Tuskegee syphilis study and the various Atomic Energy Commission radiation experiments. While recent laws try to better preserve the concepts of informed consent and minimized risk, the threat of being experimented on against one’s will still remains in the American system.

Stay tuned for some extracts from the ACHRE Final Report.

DU Armor

December 26, 2010

One of the most interesting modifications of the M1A1 series was the new armor composite including depleted uranium (DU) plate. This armor greatly increased resistance against kinetic energy rounds. During the Gulf War, M1A1 tanks could directly engage enemy tanks while in the enemy’s line-of-sight with little risk from any eventual damage from incoming retaliatory fire. This means that M1A1 tanks could hit their targets, while Iraqi tanks couldn’t hit, or, if they hit, couldn’t damage M1A1 tanks. Also, due to DU armor, not a single US tank was penetrated from enemy fire. US tanks took many close direct hits from Iraqi Soviet-made T-72 and T-72M tanks, but enemy rounds were simply not able to penetrate the M1A1 tank’s armor. The model that had this feature was called M1A1 HA (Heavy Armor), and had a protection equivalent to 600 mm against kinetic energy ammunition (APFSDS), and 1,300 mm against chemical energy warheads (ATGM’s and HEAT ammunition).

The armor protection of today’s M1A1 Abrams models is much better than that of the original M1A1 HA tanks that saw combat during the Gulf War (1991).

Fact is, a crew inside a DU armored tank is more secure than a crew inside a tank without DU armor.

Other uses of DU include : shielding containers for gamma sources, aircraft counterweights, and those vectors of radioactive dust, DU ammunition.

Congo Uranium and the Tragedy of Hiroshima

December 26, 2010

From:

http://www.fredsakademiet.dk/library/uran.pdf
(read in conjunction with:
http://www.anngarrison.com/component/content/article/24/27-congo-uranium-mining-and-the-tragedy-of-hiroshima and
http://www.friendsofthecongo.org/ )

3.8 (Fleckner)
55th Pugwash Conference
Hiroshima, Japan 22/27 July 2005
Congo Uranium and the Tragedy of Hiroshima
Mads Fleckner and John Avery
University of Copenhagen, July 2005
Congo uranium starts the Manhattan Project
The nuclear bombs that destroyed Hiroshima and Nagasaki led to the tragic deaths of a large proportion of the men, women and children living there, and the anti-human technology that obliterated the two cities still casts a very dark shadow over the future of humankind. One of the little-known aspects of this tragedy is the role played by the uranium mines of Congo. In this paper, we will review the role that Congo Uranium played in starting the Manhattan Project, and how Congo uranium was used to make the first nuclear bombs. We will also look at what is happening today at the officially closed but very active Congo uranium mines. Finally, we will examine some problems of uranium and nuclear proliferation.
In the summer of 1939, while Hitler was preparing to invade Poland, alarming news reached physicists in the United States. In addition to articles on uranium fission published in Naturwissenschaften and Deutsche Allgemeine Zeitung, two meetings of German atomic scientists had been held in Berlin under the auspices of the Research Division of the German Army Weapons Department. Furthermore, Germany had stopped the sale of uranium from mines in Czechoslovakia.
The world’s most abundant supply of uranium, however, was not in Czechoslovakia, but in Belgian Congo. Leo Szilard, a refugee Hungarian physicist living in the US, was deeply worried that the Nazis were about to construct atomic bombs; and it occurred to him that uranium from Belgian Congo should not be allowed to fall into their hands
Einstein’s fateful letter
Szilard knew that his former teacher, Albert Einstein, was a personal friend of Elisabeth, the Belgian Queen Mother. Einstein had met Queen Elisabeth and King Albert of Belgium at the Solvay Conferences, and mutual love of music had cemented a friendship between them. When Hitler came to power in 1933, Einstein had moved to the Institute of Advanced Studies at Princeton; and Szilard decided to visit him there. Szilard reasoned that because of Einstein’s great prestige,

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and because of his long-standing friendship with the Belgian Royal Family, he would be the proper person to warn the Belgians not to let their uranium fall into the hands of the Nazis.
It turned out that Einstein was vacationing at Preconic, Long Island, where he had rented a small house from a friend named Dr. Moore. Leo Szilard set out for Peconic, accompanied by the theoretical physicist, Eugene Wigner, who, like Szilard, was a Hungarian and a refugee from Hitler’s Europe.
For some time, the men drove around Peconic, unable to find Dr. Moors house. Finally Szilard, with his gift for foreseeing the future, exclaimed: “Lets give it up and go home. Perhaps fate never intended it. We should probably be making a frightful mistake in applying to any public authorities in a matter like this. Once a government gets hold of something, it never lets go”. However, Wigner insisted that it was their duty to contact Einstein and to warn the Belgians, since they might thus prevent a world catastrophe. Finally they found the house by asking a small boy in the street if he knew where Einstein lived.
Einstein agreed to write a letter to the Belgians warning them not to let uranium from the Congo into the hands of the Nazis. Wigner suggested that the American State Department ought to be notified that such a letter was being written.
On August 2, 1939, Szilard again visited Einstein, this time accompanied by Edward Teller, who (like Szilard and Wigner) was a refugee Hungarian physicist. By this time, Szilards plans had grown more ambitious; and he carried with him the draft of a letter to the American President, Franklin D. Roosevelt. Einstein made a few corrections, and then signed the fateful letter, which reads (in part) as follows:
“Some recent works of E. Fermi and L. Szilard, which has been communicated to me in manuscript, leads me to expect that the element uranium may be turned into an important source of energy in the immediate future. Certain aspects of the situation seem to call for watchfulness and, if necessary, quick action on the part of the Administration. I believe, therefore, that it is my duty to bring to your attention the following…”
It is conceivable that extremely powerful bombs of a new type may be constructed. A single bomb of this type, carried by boat and exploded in a port, might very well destroy the whole port, together with some of the surrounding territory…”
“I understand that Germany has actually stopped the sale of uranium from Czechoslovakian mines which she has taken over. That she should have taken such an early action might perhaps be understood on the ground that the son of the German Under-Secretary of State, von Weizäcker, is attached to the Kaiser Wilhelm Institute in Berlin, where some of the American work is being repeated.”
On October 11, 1939, three weeks after the defeat of Poland, Roosevelt’s economic adviser, Alexander Sachs, personally delivered the letter to the President. After discussing it with Sachs, the President commented, “This calls for action.” Later, when atomic bombs where dropped on civilian populations in an already virtually defeated Japan, Einstein bitterly regretted having signed the letter to Roosevelt. 80% of the uranium later used in the Manhattan project came from the Shinkolobwe deposit in Belgian Congo.
Shinkolobwe uranium and the collapsed state
When the Belgians left the Congo in 1960, they closed the mine by flooding the shafts and placing a concrete slab over the entrance. However, early in 2004, Arnaud Zajtman of the BBC found 6000

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people working the mine illegally. Ore from the Shinkolobwe mine is taken to smelters both in Congo and in nearby Zimbabwe.
“They are digging as fast as they can dig, and everyone is buying it,” commented John Skinner, a mining engineer based in the nearby town of Likasi. “The problem is that nobody knows where it is going. There is no control at all.”
Besides containing uranium, ore from the Shinkolobwe mine also contains a high percentage of cobalt, which is used in mobile telephones. It is this richness in cobalt that makes mining the ore especially profitable, so that motivation for closing Shinkolobwe may be lacking. Congo’s President, Joseph Kabila ordered the mine closed in March 2004, but the order has by no means been enforced. Thus Congo’s uranium not only initiated the series of events that led to the development of nuclear weapons and the tragedies of Hiroshima and Nagasaki; it also contributes to the present danger of nuclear proliferation and nuclear terrorism.
The Congo Crisis
To understand the difficulty of closing Shinkolobwe, we need to look in detail at the recent history of the Congo. The Shinkolobwe mine is located in the province of Katanga (named Shaba during the Zairian time, 1971-1997) in South-Eastern Congo. The region is, like many other regions in the Democratic Republic of Congo, rich of minerals to an extreme extent. The region’s mining, which has been influenced by Zimbabwe (Robert Mugabe was a close ally and business partner to Laurent Kabila), played a crucial role in the time of decolonization and the change of regime that led to Joseph Mobutu’s thirty two years of kleptocracy and misrule of the country.
The Congo gained independence from Belgium on June 30, 1960. The following period was politically tumultuous and resulted in a complete breakdown of law and order in the entire country. Several riots occurred as well as mutiny in the Congolese national army. On June 7 Belgium decided to reinforce its troops that remained at key bases in Congo, not only to restore law and order but also to protect the remaining Europeans. On July 10, the central government in Leopoldville (Kinshasa) asked the UN for military assistance. The Belgian colonial power had left the indigenous Congolese poorly educated with little chance to govern the country successfully. When the Katanga province declared its independence from the central government July 11, the UN had its hands full. The cold war was at its height and Congo had become a strategic playground on the African continent. That was the Congo crisis.
The UN Secretary-General, Dag Hammarskjöld, was the key architect of the Congo mission at that time. It was named ONUC (French initials), and made three attempts to prevent Moise Tshombé and his army of mercenaries from seceding the Katanga province from the rest of the Congo. The main tasks were to assist the central government in restoring law and order, and to maintain the territorial integrity of the country. Dag Hammarskjöld tragically lost his life under blurred circumstances. His plane crashed on a flight from Leopoldville (now Kinshasa) to the Katanga province. He was supposed to have met with Moise Tshombé to negotiate, but the plane never reached its destination.
Cursed by riches
The present conflict in the Democratic Republic of Congo has often been referred to as a resource based conflict, and as one crucial rapport from 2002 by the British government has put it: “…the country is cursed by riches.” The former Danish EU-commissioner, Poul Nielson actually received

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a letter about the present problem with the Shinkolobwe mine and some stored cans with radioactive material in a nearby city. But he stated that: “it is a problem for the present central transitional government to take care of,” and that, “it is the lack of a strong central government and its lack of governance in the region that is the core problem.” To some extent he is right. It is not the valuable minerals that create the problems, but the failure of politics. Poul Nielson stated that the EU couldn’t go into specific troubled cases in DR Congo or take special action in the single case, but should rather support the Congolese people and transitional government in implementing democracy and a system of laws that will provide the necessary security in dealing with Congo’s various minerals.
Besides being a mineral resource based conflict the present problems with which DR Congo struggles with are of multilayered character. Violent factionalism and ethnic affiliation are all crucial elements that fuel the ongoing conflict. Strong financial interests from at least six neighbouring countries furthermore inflame the internal struggles that create outspread instability and millions of internally displaced people.
As René Lemarchand so eloquent says it: “In its most recent avatar – the Democratic Republic of Congo – the former Belgian colony is not just a failed state; it is the epitome of the collapsed state, whose descent into hell has set loose a congeries of rival factions fighting proxy wars on behalf of half a dozen African states.”
The peace initiatives never succeeded in having the expected impact in the eastern zone of DR Congo. Various rebel groups related to or supported by the neighbouring countries have found a good business in operating in an environment that on the surface seems anarchistic. The central government in Kinshasa on the other side of the country, 1500 kilometres away, has only modest control of the situation in the east. That makes the rebel groups’ basis for negotiating better than the central government’s, with arguments and unrealistic demands motivated from not wanting peace.
The conflict in DR Congo carries elements of civil war, ethnic disputes, and regular war because of the involvement of neighbouring countries such as Uganda, Rwanda and Burundi, but also Zimbabwe, Namibia, Angola who are interfering in the conflict as a part of the Kinshasa Governments faction.
The unstable environment, which is fuelled by the lack of a strong central government, allows for the continuance of organized crime by local actors, but also from business partners and companies outside the country. A list of companies that have profited from the conflict and avoided paying high taxes, underlines the commercial benefits and possibilities for making a good deal out of the chaotic situation and creates a picture of a multilayered and multi-dynamic conflict, which is hosting actors who prefer a “Cash-in a Suitcase- economy”. Ethnic affiliation is a key element in the game of extracting valuable minerals from the country. As Koen Vlassenroot says, “At a time when the existing economic, administrative, and social patterns that have defined the local space become increasingly unstable, subject to external penetration, and unable to offer clear contexts within which people on the ground can make daily and life-choices, ethnicity indeed easily becomes an excuse for political action and violence”
Uranium, nuclear proliferation, and nuclear terrorism
The dangers presented by Congo uranium are the typical of the general link between uranium and the proliferation of nuclear weapons. One of the troubles of preventing proliferation is that civilian nuclear energy and research facilities can be, and have been, misused to produce fissionable material and bombs.

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Uranium contains several isotopes, i.e. it consists of atoms that have the same number of protons and electrons (and hence the same chemical properties) but which differ in the number of neutrons that their nuclei contain. The rare isotope U-235 is very slightly lighter than the common isotope, U-238. If uranium is to be used in nuclear weapons, the percentage of U-235 must be raised to at least 20%. (In natural uranium, the percentage of U-235 is only 0.71 %.) Since the chemical properties of U-235 and U-238 are identical, such an enrichment process depends on physical processes making use of the slight difference in mass. For example, a high-speed ultracentrifuge can be used to separate the two isotopes.
The problem in distinguishing between civil and military nuclear programs is that reactors used for generating power usually use fuel rods made of low enriched uranium (LEW), where the percentage of U-235 is 3-5%. However, if the same ultracentrifuges used to make LEW are run a little longer, they are perfectly capable of producing weapons-grade uranium. Thus it is practically impossible to distinguish between a civil nuclear program, aimed at producing electrical power, and a nuclear weapons program. The problem of making such a distinction is increased by chaotic political conditions, such as those within present-day Congo, just described.
During the next few decades, we are likely to witness a steady increase in the price of oil. Petroleum experts, such as Collin Campbell, estimate that the Hubbert Peak for oil (i.e. the year during which production and consumption reach a maximum and thereafter begin to decline) will occur within about a decade. Faced with the resulting energy crisis, many people will respond by suggesting nuclear energy as a universal answer. But given the near-impossibility of distinguishing between civil and military nuclear programs, can we risk the dangers that will result from an extremely widespread use of nuclear energy?
In 1945, the year of the tragic bombing of Hiroshima and Nagasaki, the Nobel-laureate physicist James Franck headed a committee of scientists at the University of Chicago that desperately tried to prevent the use of the bomb and also earnestly proposed ways to prevent nuclear weapons from endangering human civilization. The committee stated in its report, that the best way to stop the spread of nuclear weapons would be to prohibit the mining of uranium. This would mean forgoing the benefits of nuclear power, but the price would not be too high to pay to save humankind from the grave dangers of nuclear war. Today, 60 years later, we can see the wisdom of this recommendation of the Franck Report. Can we hope to rid the world of nuclear weapons while uranium continues to be mined? Can we rid the world of nuclear weapons while nuclear power is proposed as a universal solution to the energy crisis that will come with the rising price of petroleum? Do not uranium and nationalism, human greed and fallibility form too dangerous a mixture to be tolerated?
DR Congo serves as a key example of the dangers presented by the mixture of state implosion and proliferation of weapons material in general. One example from DR Congo gives an idea about the problem with civilian nuclear energy production and outdated facilities. When buying the uranium from Shinkolobwe, USA made a deal with Belgium. In return of the low price for the uranium USA supported Belgium in funding its peacetime nuclear energy programme. In the tumultuous time just before independence, Luc Gillon, a Belgian priest who had studied nuclear physics, imported a 50-kilowatt training and research reactor for isotope production, in the belief that the Belgian colony in Africa that helped ending the Second World War deserved its own atomic reactor. It was installed on the university campus in Kinshasa. The campus as well as the capital is built on fragile sandy ground, and the risk of erosion and landslides was and is a close reality. The forty-year-old reactor that has survived several riots and army mutinies was last upgraded in 1970. Now it is left without regularly checks and the water used to cool the fuel rods is grubby and impure. There is a serious risk of dangerous contamination from the dilapidated reactor because of the corrosion of the uranium rods. During the time under Mobutu, one of the fuel rods from this

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research reactor disappeared and was gone for twenty years. In the late nineties the Italian police succeeded getting hands on it; in blurred circumstances it had ended up in the hands of the Sicilian mafia. Another fuel rod from the same installation has totally disappeared and has still not been found. Congo became the first African member of IAEA.
Threat of spread, global control is needed
Not only the DR Congo, but also the Russian Federation, with its present lack of security in controlling nuclear materials and highly enriched uranium, can give us a glimpse of what is likely to become a true reality in the near future, namely uncontrolled spread of weapons and highly dangerous material leading to terrorist attacks against civilians with nuclear weapons. The security issue in DR Congo as well as in the Russian Federation is of multilayered character since the environment of corruption, poverty and unstable political structures creates the core of the problem – profitable proliferation. While in DR Congo the situation is a web of disorder and chaos, in the former Soviet Republic the problem are instead very poorly guarded laboratories using highly enriched uranium, in the politically fragile satellite states.
There is no lack of states or groups in the world that would like the prestigious ownership of an atomic bomb. States like Pakistan and Iran are possible purchasers of weapons grade uranium and nuclear weapons, and in the complexity of world politics today the risk of a nuclear bomb (or just a so called dirty bomb) being used is a realistic scenario. Unless the nuclear weapon states begin to take concrete steps towards the complete nuclear disarmament (in accordance with Article VI of the Non-Proliferation Treaty), and unless civil nuclear programs come under much stricter international supervision, we may well witness the explosion of a smuggled terrorist nuclear bomb in one of the world’s major cities.

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The Conflicted Responsibilities of Uranium Management and Use

December 26, 2010

In previous posts, I have shown the tension between the use of uranium and the mandated need to clean up contaminated sites and to control the substance.

I have posted the texts of publically released memos showing on the one hand the efforts to clean up sites such as the gunnery range at Fallon, USA, and the military “requirement” to fire DU shells at that and other ranges. I have shown the US CDC’s extensive health survey of Fallon which failed to find the cause of the protracted cancer cluster in that area.

I now post an extract from my booklet which relates the story of Prof. Shimizu of Kyoto University, his August 1945 radiological survey of Hiroshima and in particular that extract which deals with the military use of toxic and radiological substances in the modern world. While the US Governments (and other governments- the US being the only one open enough to allow awareness of the documents (as this is the case, the US is to be applauded, not condemned, for its relative openness).

THE FATE OF JAPANESE BIOLOGICAL, CHEMICAL AND NUCLEAR
RESEARCHERS UNDER THE US COMMAND
In exchange for information relating to Japan’s biological and chemical warfare data held by Unit 731, Lt. Col. Ryoichi Naito of that Unit, was granted war crimes immunity by the United States. He was appointed head of Green Cross, one of Japan’s top pharmaceutical companies. Other Unit 731 leaders joined him there. (Source: “Japan’s Germ Warfare and the Korean War” Lee Wha Rang, July 27, 1996. See also Dick Russell’s book, “The Man Who Knew Too Much”.) China remains sensitive to the continued under reporting of such events.
Shimizu, a member of the defunct F-go naval atomic bomb project, though innocent of any crime, was vulnerable to the dictates of the Occupation Authorities. As previously noted, the US confiscated the Hiroshima survey notes and other evidence. No doubt Shimizu and others had to remain compliant to information control and political sensitivities between Japan and the US in order to be employable in their positions.
Professor Shimizu carried out a distinguished career. He achieved “Full Professor” in the Institute of Chemical Research, Kyoto University, and in 1952 he was appointed Director of the Laboratory of Nuclear Radiation (now, the Laboratory of Atomic and Nuclear Physics, Division of States and Structures). At the same time, beginning in 1953, he served as a Lecturer in Physics at the Konan University in Kobe, Japan.” (Source: OBITUARIES, Professor Dr Sakae Shimizu (July 18, 1915 to December 13, 2003)
(Adapted by John Hubbell from obituary provided by Professor Yasuhito Isozumi, Kyoto
University, also from a bio sketch by Professor Megumi Tashiro at the time of Professor. Shimizu’s retirement to emeritus status in 1979.”)
http://www.canberra.edu.au/irps/Archives/vol17no3/obituaries.html)
The National Interest of the US determined many outcomes in the Post War era. Many
documents relating to fallout and radiation remained classified until President Clinton’s
Executive Order of the 1990s. (The DOE Openness Project). Such considerations may
account for the lack of alpha data in Shimizu’s paper.
In November 10th 1999 the US Senate debated the “Japanese Imperial Army Disclosure
Act of 1999”. In the course of debate Mrs. Feinstein stated in part : “This legislation will require the disclosure under the Freedom of Information Act classified records and
documents in the possession of the U.S. Government regarding chemical and biological experiments carried out by Japan during the course of the Second World War……. This
legislation is needed because although the Second World War ended over fifty years ago–and
with it Japan’s chemical and biological weapons experimentation programs–many of the
records and documents regarding Japan’s wartime activities remain classified and hidden in
U.S. Government archives and repositories. Even worse, according to some scholars, some of
these records are now being inadvertently destroyed… The world is entitled to a full and
complete record of what did transpire…… Professor Sheldon’s letter goes on to discuss three
examples of the destruction of documents relating to chemical and biological warfare
experiments that he is aware of: At Dugway Proving Grounds in Utah, at Fort Detrick in
Maryland, and at the Pentagon.” (Source: Congressional Record: November 10, 1999
(Senate) Page S14533-S14571 STATEMENTS ON INTRODUCED BILLS AND JOINT
RESOLUTIONS By Mrs. FEINSTEIN: S. 1902).
The United States utilized the Japanese Biological and Chemical weapons data. It kept that
data secret. It developed and built its own arsenal of similar weapons. This arsenal
complimented the US nuclear arsenal.
Likewise, the 1945 Japanese Hiroshima survey data was confiscated, kept secret for many
years and was slowly disclosed and returned.
Those Japanese personnel involved the Biological and Chemical weapons programmes and
who complied with US terms were rewarded with careers and in cash. The cash amounted to
40 million yen in current value. A US document states “data on human experiments may
prove invaluable” and said the information was “only obtainable through the skilful,
psychological approach to top-flight pathologists” involved in Unit 731 experiments.”
(Source: Australian Broadcasting Commission media report : “US paid for Japanese
human germ warfare data” Monday, August 15, 2005. Report cites the findings of
researcher Keiichi Tsuneishi, Professor at Kanagawa University).
The silence of the Japanese early surveys was easier to assure. Confiscation of data, and
censorship of Japanese reporting.
This may explain the nature and contents of Professor Shimizu’s Paper: its lack of reporting
of findings of fission product and uranium by the team he was a part of, and the vastly
different result obtained by modern Japanese researchers studying samples returned by the
US since the 1970s.
IMPLICATIONS FOR ATOMIC BOMB SURVIVIORS AND ALLIED PERSONNEL
WHO SERVED IN RADIATION AFFECTED AREAS OF JAPAN
The Australian Department of Veterans Affairs states:
“EXPOSURE TO RADIATION
15.21 The main body of Australian troops arrived in Japan on 21 February 1946,
some six months after the atomic bombing of Hiroshima and Nagasaki. DVA’s
long-standing position is that the level of radiation had fallen to acceptable
levels by the time the Australian BCOF contingent arrived in Japan. This
stance is based on advice from the Australian Radiation Protection and Nuclear
Safety Agency and its predecessors.” (Source: Report of the Review of
Veterans’ Entitlements, Chapter 15 British Commonwealth Occupation Forces
In Japan” , Department of Veterans Affairs, Australia.
http://www.veteransreview.gov.au/report/chapters/ch15.htm
The salient question this policy stance raises is: “What is an “acceptable level” of radiation exposure? In terms of internalised radionuclides of high activity but of minute size (that is, Hot Particles), many consider even level equivalent to “background” external levels unacceptable.
From the 1970s to the current era, the identification and measurement of Cesium 137 and other fission products, as well as the unfissioned fuel debris of Plutonium (Nagasaki) and Uranium (Hiroshima) have been made by Japanese scientists using returned 1945 soil samples.
Cesium 137 has a half life of 30.17 years, Uranium has a half life of millions of years,
Plutonium has a half life of 24,000 years. (Source: CDC Radiation Emergencies Radioisotope Brief Emergency Preparedness & Response).
The list of fission products present in the radiation affected of Japan in which Australian troops were stationed from February 1946 include many more long lived fission products.
Governments continue to dismiss the claims of service personnel suffering illnesses
recognised as having radiogenic origins. The dismissals are conducted on the basis that no
one who served in the affected areas of Japan internalised sufficient alpha and beta emitting particles to affect their health. That is a convenient value judgement in my view.
Gamma radiation is in the popular mind the type of radiation to be most feared.
In fact, it is the short range alpha and beta radiation which, if internalised into the body, cause the greatest damage to tissue. “If alpha emitting elements are taken into the body by inhalation or ingestion or from open wounds, serious problems such as cancer may develop.”
(Source: “Medical Response to Terrorism Preparedness and Clinical Practice,” Editor in
Chief Daniel C Keyes Editors Johnathan L Burstein Richard B Schwarz Raymond E.
Swienton, Chapter 13 Types of Radiation Explained for the Nonphysicist by Greene
Shepherd Page 138).
As for particles which emit Beta radiation: “This radiation becomes dangerous if the radioactive material is ingested or inhaled. Some of the isotopes are retained in the body for a long period in the body – sometimes permanently – when the alpha or beta radiations can be extremely dangerous, since vital organs are exposed and a cancer hazard set up.” (Source:
Alexander, Peter, “Atomic Radiation and Life”, Pelican Books, A399, 1957 Page 27)
As this hazard is based upon chance inhalation of very small radioactive particles, external monitoring of surfaces may be an inadequate guide as to safety. For example : “…the
problem that highly active particles may be present in the air although the external dose rate is below the recommended operative action level is not only theoretical.” (Source: Pollanen,
R. “Nuclear Fuel Particles in the Environment – Characteristics, Atmospheric
Transport and Skin Doses”, STUK – Radiation and Nuclear Safety Authority, University of Helsinki, Department of Physics, Academic Dissertation, presented May 28, 2002.
ISBN 951-712-528-3).

No amount of 2 dimensional monitoring can prevent the internalisation of small highly
dangerous particles which invisibly infect the 3 dimensional biosphere.
And so the innocent continue to suffer, though the terminology has changed from “fallout” to“Hot Particle”.
IMPLICATIONS FOR THE MODERN WORLD
The first implication is the severing of radiological safety knowledge. For example: “In
1943, when the first reasonable large-scale fission yields were examined by us in Berkeley,(Seaborg and MacMillan), and it was clear that fission products would be a very greathazard if we ever succeeded in obtaining chain reactions of military significance, the reportwas labeled with the reddest of classifications. I was so horrified by the biologicalconclusions (radiation hazard) of John Lawrence and Dr Stone…. In the end, if we peg away, truth will out. Let’s not jump to the lions but gird up our loins! “.(Source: – Sir Mark Oliphant, writing to Hedley Marston, 12.9.56 (AAS – Marston; 21), cited by Dr Roger Cross,
“Fallout – HedleyMarston and the British Bomb Tests in Australia,” R. Cross,
Wakefield Press Copyright Roger Cross, 2001, ISBN 1 86254 523 5.
Also severed was prewar research into nuclear medicine. If the research involved radioactivesubstances which were identified by the Manhattan Project as relevant to either atomic fallout
or neutron induced beta emitters, the research was suppressed. (Source: Brucer, Marshall, “A Chronology of Nuclear Medicine”. Heritage Publishers, St. Louis, 1990. Esp: “Chronologyfrom 1940 to 1953 – Vignettes on Manhattan District Days and Atomic Medicine.” p 259,
“The initial declassification of MED reports (Manhattan Engineering District DeclassifiedDocuments)”.
A third implication was the necessity to avoid international condemnation for the dropping ofthe atomic bomb. Such a consequence for the United States of America should the facts of
the weapon effects become fully appreciated by the world at the time were explained toSecretary of War Stimson of the USA on 11 June 1945 in a report authored by an expertgroup comprised of 3 physicists, 3 chemists and a biologist. The report was entitled “Socialand Political Implications of Atomic Energy” by Drs James Franck, Donald Hughes, Leo Szilard, Thorfin Hogness, Glenn Seaborg, Eugene Rabinowitch and J.J. Nickson
A fourth implication has been the denial of the chemical nature of nuclear and radiological weapons. This consists of 1. Direct insult to DNA 2. THE FORMATION OF ABNORMAL CHEMICALS AND CHEMICAL REACTIONS WITHIN HUMAN CELLS. Ie Ionisingradiation causes water in cells to produce hydrogen peroxide and free radicals of oxygen and hydrogen. THESE CHEMICAL PROCESSES ARE A PRIMARY CAUSE OF RADIATION SICKNESS, A CONDITION ORIGINALLY CALLED RADIATION POISONING.
(Source: Alexander, Peter, “Atomic Radiation and Life”, Pelican Books, A399, 1957,
Chapter 8 “Enter the Chemist”, pp179 – 193, in particular, page 188, “The Oxygen Effect”. Alexander’s work cites sources such as Gray, 1946, Lea, 1947, Evans, 1952, and Pollard 1954.)
A fifth implication is the downplaying of alpha and beta emitting particles in the environment as internal hazards. For example the United States and other countries maintain that Depleted Uranium ammunition is safe after use, that its dust poses no threat.

However: “Since the Department of Interior will retain the ultimate land management for all of the public lands encompassing the Range, they, as well as the Air Force should be consulted concerning the proposal to potentially disperse more than 1.5 tons of Depleted Uranium (DU) and up to 100 pounds of Beryllium on the public lands encompassing the Range.” (Source: Paul J. Liebendorfer, P.E. Bureau of Federal Facilities, State of Nevada,
Department of Conservation and Natural Resources Division of Environmental Protection 333 W. Nye Lane, Room 138 Carson City, Nevada 89706-0851, Letter dated July 12 1999 to Mr. George Laskar Assistant Area Manager Department of Energy, Albuquerque Operations office P. O. Box 5400 Albuquerque, NM 87185 http://ndep.nv.gov/boff/ndep10.htm)
And:
“Soils Project
Soils Project analyzes contaminated surface and shallow subsurface soils on the Nevada Test Site and the Nellis Test and Training Range, including the Tonopah Test Range.
Contamination at these sites is the result of historic nuclear detonations, weapons safety experiments, rocket engine development, and hydronuclear tests.
The contaminants of concern are primarily americium, plutonium, depleted uranium, and other man-made radioactive materials. In addition, there are sites where metals may be present above regulatory limits. The U.S. Department of Energy Nevada Site Office is working closely with the U.S. Air Force and the State of Nevada to determine what corrective actions may be necessary.” (Source: US Department of Energy, National Nuclear Security Administration, U.S. DOE/NNSA – Nevada Site Office Environment Management
.http://www.nv.doe.gov/emprograms/environment/restoration/soils.htm).
CONCLUSION
The nature of the atomic bomb and its fallout products were well known to the Manhattan Project by 1943.
These hazards included radioactive isotopes that while posing either no or lesser harms when outside of the body, presented grave threats to the health and safety of individuals when internalised.
Internal hazards were minimized.

The control of Japanese information regarding biological, chemical and nuclear weapons was enforced by the United States for many years. This control enabled the progress of the US biological, chemical and nuclear weapon development in secret, while in fact the program owed much to the information gained by the US from the Japanese biological and chemical weapon programmes.
This secrecy has delayed independent research and possibly forced compliance on Japanese research up until the time Professor Shimizu wrote his Paper.

In 1999 the Japanese Imperial Army Disclosure Act may have fostered more openness in researching these inter wined topics.
Hiroshima may have been a chosen target for atomic attack in part because a Head Quarters Unit of the Japanese Biological and Chemical Weapons programme was sited there, along with a factory which manufactured these munitions.

The use of Depleted Uranium weaponry continues. The intensely dangerous inhalant sized fine particles of pure uranium produced by these weapons upon firing pose a threat to health.
This threat and the need to clean up DU affected areas of the USA continues to be a matter of debate between levels and sections of the US Government. For example correspondence cited reveals Depleted Uranium and other man made radioactive substances may be above regulatory limits but that “the U.S. Department of Energy Nevada Site Office is working closely with the U.S. Air Force and the State of Nevada to determine what corrective actions may be necessary.”

The desire to continue Depleted Uranium use as a weapon is presently being balanced against regulations in US Law which recognise alpha emitters as a hazard. However modern battle tactics determine that DU weapons are fired. There is little basis for public confidence in the United States Air Force negotiating with civil safety authorities in regard to how much DU munitions debris it is able to deposit on test ranges.
Within this conflicted scenario, the victims of nuclear and radiological warfare are denied recognition, and radioactive ammunition is still fired in foreign fields of combat with no such official debate about clean up .
Professor Shimizu’s statement that “No acceptable Alpha was found”, written in 1982 about Hiroshima in August 1945, is prescient. The secret weapons programmes of both Japan and the United States have congealed into a moral dilemma. The denial of lingering harms from alpha emitting refined Uranium dust directly links to the original denial of similar harms of internalised alpha radiation in Hiroshima and Nagasaki.
The vector for affliction suffered by the pre war dial painters – procedures for the
internalisation of alpha emitters – has in effect been weaponised in various forms.
The solution was identified by Professor Shimizu in 1982. It is up to us to gain a new sense of ourselves and of the place of humanity within a higher context. There is light and dark in all nations.
Paul Langley
9 April 2008

The conficted scenario has since given rise to the mandate apparently given to Bobby Scott of Lovelace Institute New Mexico, a contractor for US DOE and DOD, who, a cursory search reveals, has travelled the world in concert with the awarding of DOE research contracts to places such as Pam Sykes, Flinders University South Australia, and other places in an attempt to over ride the mandate to control and clean up contaminated sites. In my opinion the experiments and flood of publications are an attempt to justify the use of DU weapons by magically redining uranium as a safe and beneficial by describing low level radiaiton as a public health benefit.

Alpha is high LET and an internal hazard. I have shown genetic variation produces variations in health effects. The experiments DOE funds and Scott promlugates do not hold.

Happy New Year

Toxicological Profile of Uranium – US Agency for Toxic Substances and Disease Registry

December 26, 2010

In a previous post I put up the link to the ATDSR tox. profile for Uranium, saying I might post the contents.

Here’s the contents. 400 odd pages. It would be better to download them yourself from the site. However, if Los Alamos, Bobby Scott, Lovelace Institute, Kevin Foley, Mike Rann, Flinders University have their way, they will take it down. So here is a rough as bags cut and paste.

TOXICOLOGICAL PROFILE FORURANIUM
U.S. DEPARTMENT OF HEALTH AND HUMAN SERVICES Public Health Service Agency for Toxic Substances and Disease Registry September 1999

v URANIUM FOREWORD
The Superfund Amendments and Reauthorization Act (SARA) of 1986 (Public Law 99-499) extended and amended the Comprehensive Environmental Response, Compensation, and Liability Act of 1980 (CERCLA or Superfund). This public law directed the Agency for Toxic Substances and Disease Registry (ATSDR) to prepare toxicological profiles for hazardous substances which are most commonly found at facilities on the CERCLA National Priorities List and which pose the most significant potential threat to human health, as determined by ATSDR and the Environmental Protection Agency (EPA). The lists of the 250 most significant hazardous substances were published in the Federal Register on April 17, 1987, October 20, 1988, October 26, 1989, and on October 17, 1990.
Section 104 (I) (3) of CERCLA, as amended, directs the Administrator of ATSDR to prepare a toxicological profile for each substance on the list. Each profile must include the following content:
(A) An examination, summary, and interpretation of available toxicological information and epidemiological evaluations on the hazardous substance in order to ascertain the levels of significant human exposure for the substance and the associated acute, subacute, and chronic health effects,
(B) A determination of whether adequate information on the health effects of each substance is available or in the process of development to determine levels of exposure which present a significant risk to human health of acute, subacute, and chronic health effects, and
(C) Where appropriate, an identification of toxicological testing needed to identify the types or levels of exposure that may present significant risk of adverse health effects in humans.

This toxicological profile is prepared in accordance with guidelines developed by ATSDR and EPA. The original guidelines were published in the Federal Register on April 17, 1987. Each profile will be revised and republished as necessary, but no less often than every three years, as required by SARA.
The ATSDR toxicological profile is intended to characterize succinctly the toxicological and adverse health effects information for the hazardous substance being described. Each profile identifies and reviews the key literature (that has been peer-reviewed) that describes a hazardous substance’s toxicological properties. Other pertinent literature is also presented but described in less detail than the key studies. The profile is not intended to be an exhaustive document; however, more comprehensive sources of specialty information are referenced.
Each toxicological profile begins with a public health statement, which describes in nontechnical language a substance’s relevant toxicological properties. Following the public health statement is information concerning levels of significant human exposure and, where known, significant health effects. The adequacy of information to determine a substance’s health effects is described in a health effects summary. Data needs that are of significance to protection of public health will be identified by ATSDR, the National Toxicology Program (NTP) of the Public Health Service, and EPA. The focus of the profiles is on health and toxicological information; therefore, we have included this information in the beginning of the document.

This profile reflects our assessment of all relevant toxicological testing and information that has been peer reviewed. It has been reviewed by scientists from ATSDR, the Centers for Disease Control and Prevention, the NTP, and other federal agencies. It has also been reviewed by a panel of nongovernment peer reviewers and is being made available for public review. Final responsibility for the contents and views expressed in this toxicological profile resides with ATSDR.
Jeffrey P. Koplan, M.D., M.P.H.AdministratorAgency for Toxic Substances andDisease Registry

PEER REVIEW
A peer review panel was assembled for uranium. The panel consisted of the following members:
1 Herman Cember, Ph.D., CHP, Lafayette, Indiana
2 Ron Kathren, Ph.D., CHP, Professor, Richland, Washington.
3 Paul Morrow, Ph.D., Professor Emeritus, Rochester, New York.
4 Richard Leggett, Ph.D., Research Scientist, Oak Ridge, Tennessee.
5 Darrell Fisher, Ph.D., Senior Scientist, Richland, Washington

These experts collectively have knowledge of the physical and chemical properties, toxicokinetics, key health end points, mechanisms of action, human and animal exposure, and quantification of risk to humans of uranium and uranium compounds. All reviewers were selected in conformity with the conditions for peer review specified in Section 104(i)(13) of the Comprehensive Environmental Response, Compensation, and Liability Act, as amended.
Scientists from the Agency for Toxic Substances and Disease Registry (ATSDR) have reviewed the peer reviewers’ comments and determined which comments will be included in the profile. A listing of the peer reviewers’ comments not incorporated in the profile, with a brief explanation of the rationale for their exclusion, exists as part of the administrative record for this compound. A list of databases reviewed and a list of unpublished documents cited are also included in the administrative record.
The citation of the peer review panel should not be understood to imply its approval of the profile’s final content. The responsibility for the content of this profile lies with the ATSDR.

TOXICOLOGICAL PROFILE FORURANIUM
U.S. DEPARTMENT OF HEALTH AND HUMAN SERVICESPublic Health ServiceAgency for Toxic Substances and Disease Registry
September 1999
ii URANIUM
DISCLAIMER
The use of company or product name(s) is for identification only and does not imply endorsement by the Agency for Toxic Substances and Disease Registry.
iii URANIUM
UPDATE STATEMENT
A Toxicology Profile for Uranium was released in September 1997. This edition supersedes any previously released draft or final profile.
Toxicological profiles are revised and republished as necessary, but no less than once every three years. For information regarding the update status of previously released profiles, contact ATSDR at:
Agency for Toxic Substances and Disease RegistryDivision of Toxicology/Toxicology Information Branch1600 Clifton Road NE, E-29Atlanta, Georgia 30333
.v URANIUM
FOREWORD
The Superfund Amendments and Reauthorization Act (SARA) of 1986 (Public Law 99-499) extended and amended the Comprehensive Environmental Response, Compensation, and Liability Act of 1980 (CERCLA or Superfund). This public law directed the Agency for Toxic Substances and Disease Registry (ATSDR) to prepare toxicological profiles for hazardous substances which are most commonly found at facilities on the CERCLA National Priorities List and which pose the most significant potential threat to human health, as determined by ATSDR and the Environmental Protection Agency (EPA). The lists of the 250 most significant hazardous substances were published in the Federal Register on April 17, 1987, October 20, 1988, October 26, 1989, and on October 17, 1990.
Section 104 (I) (3) of CERCLA, as amended, directs the Administrator of ATSDR to prepare a toxicological profile for each substance on the list. Each profile must include the following content:
(A) An examination, summary, and interpretation of available toxicological information and epidemiological evaluations on the hazardous substance in order to ascertain the levels of significant human exposure for the substance and the associated acute, subacute, and chronic health effects,
(B) A determination of whether adequate information on the health effects of each substance is available or in the process of development to determine levels of exposure which present a significant risk to human health of acute, subacute, and chronic health effects, and
(C) Where appropriate, an identification of toxicological testing needed to identify the types or levels of exposure that may present significant risk of adverse health effects in humans.

This toxicological profile is prepared in accordance with guidelines developed by ATSDR and EPA. The original guidelines were published in the Federal Register on April 17, 1987. Each profile will be revised and republished as necessary, but no less often than every three years, as required by SARA.
The ATSDR toxicological profile is intended to characterize succinctly the toxicological and adverse health effects information for the hazardous substance being described. Each profile identifies and reviews the key literature (that has been peer-reviewed) that describes a hazardous substance’s toxicological properties. Other pertinent literature is also presented but described in less detail than the key studies. The profile is not intended to be an exhaustive document; however, more comprehensive sources of specialty information are referenced.
Each toxicological profile begins with a public health statement, which describes in nontechnical language a substance’s relevant toxicological properties. Following the public health statement is information concerning levels of significant human exposure and, where known, significant health effects. The adequacy of information to determine a substance’s health effects is described in a health effects summary. Data needs that are of significance to protection of public health will be identified by ATSDR, the National Toxicology Program (NTP) of the Public Health Service, and EPA. The focus of the profiles is on health and toxicological information; therefore, we have included this information in the beginning of the document.
vi URANIUM
This profile reflects our assessment of all relevant toxicological testing and information that has been peer reviewed. It has been reviewed by scientists from ATSDR, the Centers for Disease Control and Prevention, the NTP, and other federal agencies. It has also been reviewed by a panel of nongovernment peer reviewers and is being made available for public review. Final responsibility for the contents and views expressed in this toxicological profile resides with ATSDR.
Jeffrey P. Koplan, M.D., M.P.H.AdministratorAgency for Toxic Substances andDisease Registry
vii URANIUM
QUICK REFERENCE FOR HEALTH CARE PROVIDERS
Toxicological Profiles are a unique compilation of toxicological information on a given hazardous substance. Each profile reflects a comprehensive and extensive evaluation, summary, and interpretation of available toxicologic and epidemiologic information on a substance. Health care providers treating patients potentially exposed to hazardous substances will find the following information helpful for fast answers to often-asked questions.
Primary Chapters/Sections of Interest
Chapter 1: Public Health Statement: The Public Health Statement can be a useful tool for educating patients about possible exposure to a hazardous substance. It explains a substance’s relevant toxicologic properties in a nontechnical, question-and-answer format, and it includes a review of the general health effects observed following exposure.
Chapter 2: Health Effects: Specific health effects of a given hazardous compound are reported by route of exposure, by type of health effect (death, systemic, immunologic, reproductive), and by length of exposure (acute, intermediate, and chronic). In addition, both human and animal studies are reported in this section. NOTE: Not all health effects reported in this section are necessarily observed in the clinical setting. Please refer to the Public Health Statement to identify general health effects observed following exposure.
Pediatrics: Four new sections have been added to each Toxicological Profile to address child health issues: Section 1.6 How Can (Chemical X) Affect Children? Section 1.7 How Can Families Reduce the Risk of Exposure to (Chemical X)? Section 2.6 Children’s Susceptibility Section 5.6 Exposures of Children
Other Sections of Interest: Section 2.7 Biomarkers of Exposure and Effect Section 2.10 Methods for Reducing Toxic Effects
ATSDR Information Center Phone: 1-888-42-ATSDR or 404-639-6357 Fax: 404-639-6359 E-mail: atsdric@cdc.gov Internet: http://atsdr1.atsdr.cdc.gov:8080
The following additional material can be ordered through the ATSDR Information Center:
Case Studies in Environmental Medicine: Taking an Exposure History—The importance of taking an exposure history and how to conduct one are described, and an example of a thorough exposure history is provided. Other case studies of interest include Reproductive and Developmental Hazards; Skin Lesions and Environmental Exposures; Cholinesterase-Inhibiting Pesticide Toxicity; and numerous chemical-specific case studies.
URANIUM viii
Managing Hazardous Materials Incidents is a three-volume set of recommendations for on-scene (prehospital) and hospital medical management of patients exposed during a hazardous materials incident. Volumes I and II are planning guides to assist first responders and hospital emergency department personnel in planning for incidents that involve hazardous materials. Volume III—Medical Management Guidelines for Acute Chemical Exposures—is a guide for health care professionals treating patients exposed to hazardous materials.
Fact Sheets (ToxFAQs) provide answers to frequently asked questions about toxic substances.
Other Agencies and Organizations
The National Center for Environmental Health (NCEH) focuses on preventing or controlling disease, injury, and disability related to the interactions between people and their environment outside the workplace. Contact: NCEH, Mailstop F-29, 4770 Buford Highway, NE, Atlanta, GA 303413724 • Phone: 770-488-7000 • FAX: 770-488-7015.
The National Institute for Occupational Safety and Health (NIOSH) conducts research on occupational diseases and injuries, responds to requests for assistance by investigating problems of health and safety in the workplace, recommends standards to the Occupational Safety and Health Administration (OSHA) and the Mine Safety and Health Administration (MSHA), and trains professionals in occupational safety and health. Contact: NIOSH, 200 Independence Avenue, SW, Washington, DC 20201 • Phone: 800-356-4674 or NIOSH Technical Information Branch, Robert A. Taft Laboratory, Mailstop C-19, 4676 Columbia Parkway, Cincinnati, OH 45226-1998
• Phone: 800-35-NIOSH.
The National Institute of Environmental Health Sciences (NIEHS) is the principal federal agency for biomedical research on the effects of chemical, physical, and biologic environmental agents on human health and well-being. Contact: NIEHS, PO Box 12233, 104 T.W. Alexander Drive, Research Triangle Park, NC 27709 • Phone: 919-541-3212.
Referrals
The Association of Occupational and Environmental Clinics (AOEC) has developed a network of clinics in the United States to provide expertise in occupational and environmental issues. Contact: AOEC, 1010 Vermont Avenue, NW, #513, Washington, DC 20005 • Phone: 202-347-4976 • FAX: 202-347-4950 • e-mail: aoec@dgs.dgsys.com • AOEC Clinic Director: http://occ-envmed.mc.duke.edu/oem/aoec.htm.
The American College of Occupational and Environmental Medicine (ACOEM) is an association of physicians and other health care providers specializing in the field of occupational and environmental medicine. Contact: ACOEM, 55 West Seegers Road, Arlington Heights, IL 60005 • Phone: 847-228-6850 • FAX: 847-228-1856.
ix URANIUM
CONTRIBUTORS
CHEMICAL MANAGER(S)/AUTHORS(S):
Sam Keith, M.S., CHPATSDR, Division of Toxicology, Atlanta, GA
Wayne Spoo, DVM, DABTResearch Triangle Institute, Research Triangle Park, NC
James Corcoran, Ph.D. Research Triangle Institute, Research Triangle Park, NC
THE PROFILE HAS UNDERGONE THE FOLLOWING ATSDR INTERNAL REVIEWS:
1 Health Effects Review. The Healths Effects Review Committee examines the health effects chapter of each profile for consistency and accuracy in interpreting health effects and classifying end points.
2 Minimal Risk Level Review. The Minimal Risk Level Workgroup considers issues relevant to substance-specific minimal risk levels (MRLs), reviews the health effects database of each profile, and makes recommendations for derivation of MRLs.

.xi URANIUM
PEER REVIEW
A peer review panel was assembled for uranium. The panel consisted of the following members:
1 Herman Cember, Ph.D., CHP, Lafayette, Indiana
2 Ron Kathren, Ph.D., CHP, Professor, Richland, Washington.
3 Paul Morrow, Ph.D., Professor Emeritus, Rochester, New York.
4 Richard Leggett, Ph.D., Research Scientist, Oak Ridge, Tennessee.
5 Darrell Fisher, Ph.D., Senior Scientist, Richland, Washington

These experts collectively have knowledge of the physical and chemical properties, toxicokinetics, key health end points, mechanisms of action, human and animal exposure, and quantification of risk to humans of uranium and uranium compounds. All reviewers were selected in conformity with the conditions for peer review specified in Section 104(i)(13) of the Comprehensive Environmental Response, Compensation, and Liability Act, as amended.
Scientists from the Agency for Toxic Substances and Disease Registry (ATSDR) have reviewed the peer reviewers’ comments and determined which comments will be included in the profile. A listing of the peer reviewers’ comments not incorporated in the profile, with a brief explanation of the rationale for their exclusion, exists as part of the administrative record for this compound. A list of databases reviewed and a list of unpublished documents cited are also included in the administrative record.
The citation of the peer review panel should not be understood to imply its approval of the profile’s final content. The responsibility for the content of this profile lies with the ATSDR.
.URANIU M xiii
CONTENTS
FOREWORD …………………………………………………………….. v
QUICK REFERENCE FOR HEALTH CARE PROVIDERS…………………………… vii
CONTRIBUTORS ………………………………………………………….ix
PEER REVIEW ……………………………………………………………xi
LIST OF FIGURES ………………………………………………………. xvii
LIST OF TABLES…………………………………………………………xix
1. PUBLIC HEALTH STATEMENT……………………………………………. 1
1.1 WHAT IS URANIUM? ………………………………………………. 1
1.2 WHAT HAPPENS TO URANIUM WHEN IT ENTERS THE ENVIRONMENT?……… 5
1.3 HOW MIGHT I BE EXPOSED TO URANIUM? …………………………….. 7
1.4 HOW CAN URANIUM ENTER AND LEAVE MY BODY? ……………………. 8
1.5 HOW CAN URANIUM AFFECT MY HEALTH? ……………………………. 9
1.6 HOW CAN URANIUM AFFECT CHILDREN? ……………………………. 10
1.7 HOW CAN FAMILIES REDUCE THE RISK OF EXPOSURE TO URANIUM? …….. 11
1.8 IS THERE A MEDICAL TEST TO DETERMINE WHETHER I HAVE BEEN EXPOSED TO URANIUM? ………………………………………….. 12
1.9 WHAT RECOMMENDATIONS HAS THE FEDERAL GOVERNMENT MADE TO PROTECT HUMAN HEALTH? ……………………………………….. 13
1.10 WHERE CAN I GET MORE INFORMATION? ……………………………. 14
2. HEALTH EFFECTS…………………………………………………….. 17
2.1 INTRODUCTION …………………………………………………. 17
2.2 DISCUSSION OF HEALTH EFFECTS BY ROUTE OF EXPOSURE …………….. 24
2.2.1 Inhalation Exposure ………………………………………….. 26
2.2.1.1 Death …………………………………………….. 27
2.2.1.2 Systemic Effects …………………………………….. 30
2.2.1.3 Immunological and Lymphoreticular Effects …………………. 81
2.2.1.4 Neurological Effects ………………………………….. 83
2.2.1.5 Reproductive Effects ………………………………….. 84
2.2.1.6 Developmental Effects ………………………………… 85
2.2.1.7 Genotoxic Effects ……………………………………. 85
2.2.1.8 Cancer ……………………………………………. 86
2.2.2 Oral Exposure ……………………………………………… 92
2.2.2.1 Death …………………………………………….. 92
2.2.2.2 Systemic Effects …………………………………….. 93
2.2.2.3 Immunological and Lymphoreticular Effects ………………… 132
2.2.2.4 Neurological Effects …………………………………. 133
2.2.2.5 Reproductive Effects …………………………………. 133
2.2.2.6 Developmental Effects ……………………………….. 135
2.2.2.7 Genotoxic Effects …………………………………… 136
2.2.2.8 Cancer …………………………………………… 137
2.2.3 Dermal Exposure …………………………………………… 138
URANIUM xiv
2.2.3.1 Death ……………………………………………. 138
2.2.3.2 Systemic Effects ……………………………………. 139
2.2.3.3 Immunological and Lymphoreticular Effects ………………… 148
2.2.3.4 Neurological Effects …………………………………. 149
2.2.3.5 Reproductive Effects …………………………………. 149
2.2.3.6 Developmental Effects ……………………………….. 149
2.2.3.7 Genotoxic Effects …………………………………… 149
2.2.3.8 Cancer …………………………………………… 149
2.3 TOXICOKINETICS……………………………………………….. 150
2.3.1 Absorption ……………………………………………….. 150
2.3.1.1 Inhalation Exposure………………………………….. 150
2.3.1.2 Oral Exposure ……………………………………… 152
2.3.1.3 Dermal Exposure……………………………………. 153
2.3.2 Distribution ………………………………………………. 154
2.3.2.1 Inhalation Exposure………………………………….. 154
2.3.2.2 Oral Exposure ……………………………………… 157
2.3.2.3 Dermal Exposure……………………………………. 159
2.3.2.4 Other Routes of Exposure ……………………………… 159
2.3.3 Metabolism……………………………………………….. 160
2.3.4 Elimination and Excretion …………………………………….. 160
2.3.4.1 Inhalation Exposure………………………………….. 160
2.3.4.2 Oral Exposure ……………………………………… 163
2.3.4.3 Dermal Exposure……………………………………. 164
2.3.5 Physiologically Based Pharmacokinetic (PBPK)/Pharmacodynamic (PD) Models 164
2.4 MECHANISMS OF ACTION ………………………………………… 181
2.4.1 PharmacokineticMechanisms ………………………………….. 181
2.4.2 Mechanisms of Toxicity………………………………………. 182
2.4.3 Animal-to-HumanExtrapolations ……………………………….. 185
2.5 RELEVANCE TO PUBLIC HEALTH ………………………………….. 186
2.6 CHILDREN’S SUSCEPTIBILITY …………………………………….. 212
2.7 BIOMARKERS OF EXPOSURE AND EFFECT …………………………… 215
2.7.1 Biomarkers Used to Identify or Quantify Exposure to Uranium …………… 216
2.7.2 Biomarkers Used to Characterize Effects Caused by Uranium ……………. 217
2.8 INTERACTIONS WITH OTHER CHEMICALS …………………………… 218
2.9 POPULATIONS THAT ARE UNUSUALLY SUSCEPTIBLE …………………. 218
2.10 METHODS FOR REDUCING TOXIC EFFECTS ………………………….. 219
2.10.1 Reducing Peak Absorption Following Exposure ……………………… 219
2.10.2 Reducing Body Burden ………………………………………. 219
2.10.3 Interfering with the Mechanism of Action for Toxic Effects……………… 220
2.11 ADEQUACY OF THE DATABASE……………………………………. 220
2.11.1 Existing Information on Health Effects of Uranium …………………… 220
2.11.2 Identification of Data Needs …………………………………… 223
2.11.3 Ongoing Studies……………………………………………. 232
3. CHEMICAL AND PHYSICAL INFORMATION……………………………….. 235
3.1 CHEMICAL IDENTITY ……………………………………………. 235
3.2 PHYSICAL, CHEMICAL, AND RADIOLOGICAL PROPERTIES ……………… 235
URANIUM xv
4. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL ………………………. 245
4.1 PRODUCTION ………………………………………………….. 245
4.2 IMPORT/EXPORT ……………………………………………….. 249
4.3 USE…………………………………………………………… 253
4.4 DISPOSAL ……………………………………………………… 253
5. POTENTIAL FOR HUMAN EXPOSURE ……………………………………. 257
5.1 OVERVIEW…………………………………………………….. 257
5.2 RELEASES TO THE ENVIRONMENT …………………………………. 262
5.2.1 Air …………………………………………………….. 262
5.2.2 Water …………………………………………………… 266
5.2.3 Soil …………………………………………………….. 267
5.3 ENVIRONMENTAL FATE………………………………………….. 270
5.3.1 Transport and Partitioning …………………………………….. 270
5.3.2 Transformation and Degradation ………………………………… 276
5.3.2.1 Air ……………………………………………… 276
5.3.2.2 Water ……………………………………………. 276
5.3.2.3 Sediment and Soil …………………………………… 277
5.4 LEVELS MONITORED OR ESTIMATED IN THE ENVIRONMENT …………… 278
5.4.1 Air …………………………………………………….. 278
5.4.2 Water …………………………………………………… 280
5.4.3 Soil …………………………………………………….. 285
5.4.4 Other Environmental Media……………………………………. 287
5.5 GENERAL POPULATION AND OCCUPATIONAL EXPOSURE ……………… 289
5.6 EXPOSURES OF CHILDREN ……………………………………….. 291
5.7 POPULATIONS WITH POTENTIALLY HIGH EXPOSURES…………………. 293
5.8 ADEQUACY OF THE DATABASE……………………………………. 294
5.8.1 Identification of Data Needs …………………………………… 294
5.8.2 Ongoing Studies……………………………………………. 297
6. ANALYTICAL METHODS………………………………………………. 299
6.1 BIOLOGICAL MATERIALS ………………………………………… 299
6.1.1 Internal Uranium Measurements ………………………………… 299
6.1.2 In Vivo andIn Vitro Uranium Measurements ……………………….. 300
6.2 ENVIRONMENTAL SAMPLES………………………………………. 304
6.2.1 Field Measurements of Uranium ………………………………… 304
6.2.2 Laboratory Analysis of Environmental Samples ……………………… 305
6.3 ADEQUACY OF THE DATABASE……………………………………. 315
6.3.1 Identification of Data Needs …………………………………… 315
6.3.2 Ongoing Studies……………………………………………. 316
7. REGULATIONS AND ADVISORIES ………………………………………. 319
8. REFERENCES ……………………………………………………….. 333
9. GLOSSARY …………………………………………………………. 381
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LIST OF APPENDICES
A. ATSDR MINIMAL RISK LEVELS AND WORKSHEETS……………………. A-1
B. USER’S GUIDE ………………………………………………….. B-1
C. ACRONYMS, ABBREVIATIONS, AND SYMBOLS ……………………….. C-1
D. OVERVIEW OF BASIC RADIATION PHYSICS, CHEMISTRY AND BIOLOGY ….. D-1
I. INDEX …………………………………………………………. I-1
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LIST OF FIGURES
2-1 Levels of Significant Exposure to Uranium, Inhalation—Chemical Toxicity ……………. 54 2-2 Levels of Significant Exposure to Uranium, Inhalation—Radiation Toxicity …………… 64 2-3 Levels of Significant Exposure to Uranium, Oral—Chemical Toxicity ………………. 115 2-4 Conceptual Representation of a Physiologically Based Pharmacokinetic (PBPK) Model for a
Hypothetical Chemical Substance ………………………………………… 1662-5 Respiratory Tract Compartments in Which Particles May Be Deposited ………………. 1682-6 Compartment Model to Represent Time-Dependent Particle Transport in the Respiratory Tract 1702-7 The Human Respiratory Tract Model: Absorption into Blood……………………… 1752-8 Biokinetic Model for Uranium After Uptake to Blood…………………………… 1772-9 Multicompartmental Model …………………………………………….. 1782-10 Existing Information on the Health Effects of Uranium………………………….. 2224-1 Flow Chart of Uranium Ore Processing …………………………………….. 2485-1 Frequency of NPL Sites with Uranium (U-238) Contamination ……………………. 2585-2 Major DOE Offices, Facilities, and Laboratories ………………………………. 2645-3 Environmental Pathways for Potential Human Health Effects from Uranium …………… 2715-4 Average Uranium Concentrations in Drinking Water for States Where Concentration
Exceeds 1 pCi/L …………………………………………………….. 2817-1 Nuclear Regulatory Commission Agreement States…………………………….. 332
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LIST OF TABLES
2-1 Levels of Significant Exposure to Uranium, Chemical Toxicity—Inhalation ……………. 31 2-2 Levels of Significant Exposure to Uranium, Radiation Toxicity—Inhalation ……………. 63 2-3 Levels of Significant Exposure to Uranium, Chemical Toxicity—Oral ………………… 94 2-4 Levels of Significant Exposure to Uranium, Chemical Toxicity—Dermal …………….. 140 2-5 Reference Respiratory Values for a General Caucasian Population at Different Levels of Activity169 2-6 Reference Values of Parameters for the Compartment Model to Represent
Time-dependent Particle Transport from the Human Respiratory Tract ………………. 1722-7 Sensitivity and Calculated Transfer Coefficients (d-1) …………………………… 1802-8 Enriched, Natural, and Depleted Uranium Mass Equivalents for Radiological Effects …….. 1932-9 Genotoxicity of Uranium In Vivo …………………………………………. 2092-10 Genotoxicity of Uranium In Vitro ………………………………………… 2102-11 Ongoing Studies on Health Effects of Uranium ……………………………….. 2333-1 Chemical Identity of Uranium Metal ………………………………………. 2363-2 Physical and Chemical Properties of Selected Uranium Compounds ………………… 2373-3 Percent Occurrence and Radioactive Properties of Naturally Occurring Isotopes of Uranium . . 2403-4 Uranium Isotope Decay Series Showing the Decay Products of the Naturally Occurring
Isotopes of Uranium ………………………………………………….. 2414-1 Uranium Ores ………………………………………………………. 2464-2 Uranium Production in the United States by Uranium Mills and Other Sources …………. 2504-3 Uranium Mining Production, 1985–1998……………………………………. 2514-4 Import of Uranium and Compounds (in kg) into the United States ………………….. 2524-5 Export of Uranium and Compounds (in kg) ………………………………….. 254
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5-1 Normalized Uranium Effluent Discharges from Uranium Mining, Milling,
Conversion, Enrichment, and Fuel Fabrication ……………………………….. 2685-2 Uranium in Airborne Particles (Composites) …………………………………. 2795-3 Uranium Analyses of Select Precipitation Composite Samples March–May 1993 ……….. 2835-4 Uranium in Rocks and Soils …………………………………………….. 2865-5 Concentrations of Uranium in Some Foods ………………………………….. 2885-6 Ongoing Studies on Environmental Effects of Uranium …………………………. 2986-1 Analytical Methods for Determining Uranium in Biological Samples ………………… 3026-2 Analytical Methods for Determining Uranium in Environmental Samples …………….. 3066-3 Additional Analytical Methods for Determining Uranium in Environmental Samples …….. 3136-4 Ongoing Studies on Analytical Methods for Uranium …………………………… 3177-1 Regulations and Guidelines Applicable to Uranium ……………………………. 324
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This public health statement tells you about uranium and the effects of exposure.
The Environmental Protection Agency (EPA) identifies the most serious hazardous waste sites in the nation. These sites make up the National Priorities List (NPL) and are the sites targeted for long-term federal cleanup activities. Elevated uranium levels have been found in at least 54 of the 1,517 current or former NPL sites. However, the total number of NPL sites evaluated for this substance is not known. As more sites are evaluated, the sites at which uranium is found may increase. This information is important because exposure to this substance may harm you and because these sites may be sources of exposure.
When a substance is released from a large area, such as an industrial plant, or from a container, such as a drum or bottle, it enters the environment. This release does not always lead to exposure. You are normally exposed to a substance only when you come in contact with it. You may be exposed by breathing, eating, or drinking the substance or by skin contact. However, since uranium is radioactive, you can also be exposed to its radiation if you are near it.
If you are exposed to uranium, many factors determine whether you’ll be harmed. These factors include the dose (how much), the duration (how long), and how you come in contact with it. You must also consider the other chemicals you’re exposed to and your age, sex, diet, family traits, lifestyle, and state of health.
1.1 WHAT IS URANIUM?
Uranium is a natural and commonly occurring radioactive element. It is found in very small amounts in nature in the form of minerals, but may be processed into a silver-colored metal. Rocks, soil, surface and underground water, air, and plants and animals all contain varying amounts of uranium. Typical concentrations in most materials are a few parts per million (ppm). This corresponds to around 4 tons of uranium in 1 square mile of soil 1 foot deep, or about half a teaspoon of uranium in a typical 8-cubic yard dump truck load of soil. Some rocks and soils may
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also contain greater amounts of uranium. If the amount is great enough, the uranium may be present in commercial quantities and can be mined. After the uranium is extracted, it is converted into uranium dioxide or other chemical forms by a series of chemical processes known as milling. The residue remaining after the uranium has been extracted is called mill tailings. Mill tailings contain a small amount of uranium, as well as other naturally radioactive waste products such as radium and thorium.
Natural uranium is a mixture of three types (or isotopes) of uranium, written as 234U, 235U, and 238U, or as U-234, U-235, and U-238, and read as uranium two thirty-four, etc. All three isotopes behave the same chemically, so any combination of the three would have the same chemical effect on your body. But they are different radioactive materials with different radioactive properties. That is why we must look at the actual percentages of the three isotopes in a sample of uranium to determine how radioactive the uranium is. For uranium that has been locked inside the earth for millions of years, we know the percentage of each isotope by weight and by radioactivity. By weight, natural uranium is about 0.01% 234U, 0.72% 235U, and 99.27% 238U. About 48.9% of the radioactivity is associated with 234U, 2.2% is associated with 235U, and 48.9% is associated with 238U.
The weight and radioactivity percentages are different because each isotope has a different physical half-life. Radioactive isotopes are constantly changing into different isotopes by giving off radiation. The half-life is the time it takes for half of that uranium isotope to give off its radiation and change into a different element. The half-lives of uranium isotopes are very long (244 thousand years for 234U, 710 million years for 235U, and 41/2 billion years for 238U). The shorter half-life makes 234U the most radioactive, and the longer half-life makes 238U the least radioactive. If you have one gram of each isotope side by side, the 234U will be about 20 thousand times more radioactive and the 235U will be 6 times more radioactive than the 238U.
Uranium is measured in units of mass (grams) or radioactivity (curies or becquerels), depending on the type of equipment available or the level that needs to be measured. The becquerel (Bq) is a new international unit, and the curie (Ci) is a traditional unit; both are currently used. A Bq is the amount of radioactive material in which 1 billion atoms transform every second, and a Ci is
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the amount of radioactive material in which 37 billion atoms transform every second. The mass and activity ratios given in the previous paragraph are those found in rocks inside the earth’s crust, where 1.5 gram of uranium is equivalent to 1 millionth of a Ci (μCi). Although this ratio can vary in air, soil, and water, the conversions made in this profile use the 1.5-to-1 ratio unless the actual isotope ratios are known. When both mass and radioactivity units are shown, the first is normally the one reported in the literature. Some of the values may be rounded to make the text easier to read.
The uranium isotopes in the earth were present when the earth was formed. Both 235U and 238U have such long half-lives that part of the uranium originally on earth is still here, waiting to give off its radiation. The original 234U would have decayed away by now, but new 234U is constantly being made from the decay of 238U. When 238U gives off its radiation, it changes or decays through a series of different radioactive materials, including 234U. This series, or decay chain, ends when a stable, non-radioactive substance is made. This element is lead. This toxicological profile deals with the uranium isotopes and not with the other radioactive decay products, like radium, thorium, and radon.
For uranium that has been in contact with water, the natural weight and radioactivity percentages can vary slightly from the percentages mentioned in the previous paragraphs. We don’t fully understand why that happens in nature, but measurements show us that it does. The processing of uranium for industrial and governmental use can also change the ratios. We give these ratios special names if they were changed by human activities. If the fraction of 235U is increased, we call it enriched uranium. However, if the portion of 235U is decreased, we call it depleted uranium. The differences between the weight and radioactivity ratios matter when we want to convert between radioactivity and mass, and when we talk about how toxic uranium might be. Depleted uranium is less radioactive than natural uranium, and enriched uranium is more radioactive than natural uranium. This profile focuses on natural and depleted uranium, which are more likely to be chemical hazards than radiation hazards. The profile also discusses enriched uranium, which can be both a chemical and a radiation hazard.
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The industrial process called enrichment is used to increase the amount of 234U and 235U and decrease the amount of 238U in natural uranium. The product of this process is enriched uranium, and the leftover is depleted uranium. Enriched uranium is more radioactive than natural uranium, and natural uranium is more radioactive than depleted uranium. When enriched uranium is 97.5% pure 235U, the same weight of enriched uranium has about 75 times the radioactivity as natural uranium. This is because enriched uranium also contains 234U, which is even more radioactive than 235U. The 235U is responsible for most of the radioactivity in enriched uranium. Natural uranium is typically about two times more radioactive than depleted uranium. Other isotopes of uranium called 232U and 233U are produced by industrial processes. These are also much more radioactive than natural uranium.
The total amount of natural uranium on earth stays almost the same because of the very long half-lives of the uranium isotopes. The natural uranium can be moved from place to place by nature or by people, and some uranium is removed from the earth by mining. When rocks are broken up by water or wind, uranium becomes a part of the soil. When it rains, the soil containing uranium can be carried into rivers and lakes. Wind can blow dust that contains uranium into the air.
Natural uranium is radioactive but poses little radioactive danger because it gives off very small amounts of radiation. Uranium transforms into another element and gives off radiation. In this way uranium transforms into thorium and gives off a particle called an alpha particle or alpha radiation. Uranium is called the parent, and thorium is called the transformation product. When the transformation product is radioactive, it keeps transforming until a stable product is formed. During these decay processes, the parent uranium, its decay products, and their subsequent decay products each release radiation. Radon and radium are two of these products. Unlike other kinds of radiation, the alpha radiation ordinarily given off by uranium cannot pass through solid objects, such as paper or human skin. For more information on radiation, see Appendix D and the glossary at the end of this profile or the ASTDR Toxicological Profile for Ionizing Radiation.
The main civilian use of uranium is in nuclear power plants and on helicopters and airplanes. It is also used by the armed forces as shielding to protect Army tanks, parts of bullets and missiles
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to help them go through enemy armored vehicles, as a source of power, and in nuclear weapons. Very small amounts are used to make some ceramic ornament glazes, light bulbs, photographic chemicals, and household products. Some fertilizers contain slightly higher amounts of natural uranium. For more information about the properties and uses of uranium, see Chapters 3, 4, and 5.
1.2 WHAT HAPPENS TO URANIUM WHEN IT ENTERS THE ENVIRONMENT?
Uranium is a naturally occurring radioactive material that is present to some degree in almost everything in our environment, including soil, rocks, water, and air. It is a reactive metal, so it is not found as free uranium in the environment. In addition to the uranium naturally found in minerals, the uranium metal and compounds that are left after humans mine and process the minerals can also be released back to the environment in mill tailings. This uranium can combine with other chemicals in the environment to form other uranium compounds. Each of these uranium compounds dissolves to its own special extent in water, ranging from not soluble to very soluble. This helps determine how easily the compound can move through the environment, as well as how toxic it might be.
The amount of uranium that has been measured in air in different parts of the United States by EPA ranges from 0.011 to 0.3 femtocuries (0.00002 to 0.00045 micrograms) per cubic meter (m3). (One femtocurie is equal to 1 picocurie [pCi] divided by 1,000. A picocurie [pCi] is 1 one-trillionth of a curie and a microgram [μg] is one millionth of a gram Even at the higher concentration, there is so little uranium in a cubic meter of air that less than one atom transforms each day.
In the air, uranium exists as dust. Very small dust-like particles of uranium in the air fall out of the air onto surface water, plant surfaces, and soil either by themselves or when rain falls. These particles of uranium eventually end up back in the soil or in the bottoms of lakes, rivers, and ponds, where they stay and mix with the natural uranium that is already there.
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Uranium in water comes from different sources. Most of it comes from dissolving uranium out of rocks and soil that water runs over and through. Only a very small part is from the settling of uranium dust out of the air. Some of the uranium is simply suspended in water, like muddy water. The amount of uranium that has been measured in drinking water in different parts of the United States by EPA is generally less than 1.5 μg (1 pCi) for every liter of water. EPA has found that the levels of uranium in water in different parts of the United States are extremely low in most cases, and that water containing normal amounts of uranium is usually safe to drink. Because of the nature of uranium, not much of it gets into fish or vegetables, and most of that which gets into livestock is eliminated quickly in urine and feces.
Uranium is found naturally in soil in amounts that vary over a wide range, but the typical concentration is 3 μg (2 pCi) per gram of soil. Additional uranium can be added by industrial activities. Soluble uranium compounds can combine with other substances in the environment to form other uranium compounds. Uranium compounds may stay in the soil for thousands of years without moving downward into groundwater. When large amounts of natural uranium are found in soil, it is usually soil with phosphate deposits. The amount of uranium that has been measured in the phosphate-rich soils of north and central Florida ranges from 4.5 to 83.4 pCi of uranium in every gram of soil. In areas like New Mexico, where uranium is mined and processed, the amount of uranium per gram of soil ranges between 0.07 and 3.4 pCi (0.1–5.1 micrograms [μg]) of uranium per gram soil). The amount of uranium in soil near a uranium fuel fabrication facility in the state of Washington ranges from 0.51 to 3.1 pCi/gram (0.8–4.6 μg uranium/gram soil), with an average value of 1.2 pCi/gram (1.7 μg uranium/gram soil). These levels must be carefully compared with the levels in uncontaminated soil in that area, since they are within the normal ranges for uncontaminated soil.
Plants can absorb uranium from the soil onto their roots without absorbing it into the body of the plant. Therefore, root vegetables like potatoes and radishes that are grown in uranium-contaminated soil may contain more uranium than if the soil contained levels of uranium that were natural for the area. Washing the vegetable or removing its skin often removes most or all of the uranium. For more information about what happens to uranium in the environment, see Chapter 5.
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1.3 HOW MIGHT I BE EXPOSED TO URANIUM?
Since uranium is found everywhere in small amounts, you always take it into your body from the air, water, food, and soil. Food and water have small amounts of natural uranium in them. People eat about 1–2 micrograms (0.6–1.0 picocuries) of natural uranium every day with their food and take in about 1.5 micrograms (0.8 picocuries) of natural uranium for every liter of water they drink, but they breathe in much lower amounts. Root vegetables, such as beets and potatoes, tend to have a bit more uranium than other foods. In a few places, there tends to be more natural uranium in the water than in the food. People in these areas naturally take in more uranium from their drinking water than from their foods. It is possible that you may eat and drink more uranium if you live in an area with naturally higher amounts of uranium in the soil or water or if you live near a uranium-contaminated hazardous waste site. You can also take in (or ingest) more uranium if you eat food grown in contaminated soil, or drink water that has unusually high levels of uranium. Normally, very little of the uranium in lakes, rivers, or oceans gets into the fish or seafood we eat. The amounts in the air are usually so small that they can be safely ignored. People who are artists, art or craft teachers, ceramic hobbyists, or glass workers who still use certain banned uranium-containing glazes or enamels may also be near to higher levels of uranium, but they will not necessarily take any into their bodies. People who work at factories that process uranium, work with phosphate fertilizers, or live near uranium mines have a chance of taking in more uranium than most other people. People who work on gyroscopes, helicopter rotor counterbalances, or control surfaces of airplanes may work with painted uranium metal, but the coating normally will keep them from taking in any uranium. People who work with armor-piercing weapons that contain uranium will be exposed to low levels of radiation while close to these weapons, but are not likely to take in any uranium. Those who fire uranium weapons, work with weapons with damaged uranium, or on equipment which has been bombarded with these weapons can be exposed to uranium and may wear protective clothes and masks to limit their intake. Larger-than-normal amounts of uranium might also enter the environment from erosion of tailings from mines and mills for uranium and other metals. Accidental discharges from uranium processing plants are possible, but these compounds spread out quickly into the air. For more information about how you may be exposed to uranium, see Chapter 5.
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1.4 HOW CAN URANIUM ENTER AND LEAVE MY BODY?
We take uranium into our bodies in the food we eat, water we drink, and air we breathe. What we take in from industrial activities is in addition to what we take in from natural sources.
When you breathe uranium dust, some of it is exhaled and some stays in your lungs. The size of the uranium dust particles and how easily they dissolve determines where in the body the uranium goes and how it leaves your body. Uranium dust may consist of small, fine particles and coarse, big particles. The big particles are caught in the nose, sinuses, and upper part of your lungs where they are blown out or pushed to the throat and swallowed. The small particles are inhaled down to the lower part of your lungs. If they do not dissolve easily, they stay there for years and cause most of the radiation dose to the lungs from uranium. They may gradually dissolve and go into your blood. If the particles do dissolve easily, they go into your blood more quickly. A small part of the uranium you swallow will also go into your blood. The blood carries uranium throughout your body. Most of it leaves in your urine in a few days, but a little stays in your kidneys and bones.
When you eat foods and drink liquids containing uranium, most of it leaves within a few days in your feces and never enters your blood. A small portion will get into your blood and will leave your body through your urine within a few days. The rest can stay in your bones, kidneys, or other soft tissues. A small amount goes to your bones and may stay there for years. Most people have a very small amounts of uranium, about 1/5,000th of the weight of an aspirin tablet, in their bodies, mainly in their bones.
Although uranium is weakly radioactive, most of the radiation it gives off cannot travel far from its source. If the uranium is outside your body, such as in soil, most of its radiation cannot penetrate your skin and enter your body. To be exposed to radiation from uranium, you have to eat, drink, or breathe it, or get it on your skin. If uranium transformation products are also present, you can be exposed to their radiation at a distance. For more information about how uranium can leave your body, see Chapter 2.
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1.5 HOW CAN URANIUM AFFECT MY HEALTH?
To protect the public from the harmful effects of toxic chemicals and to find ways to treat people who have been harmed, scientists must determine what the harmful effects are, how to test for them, how much of the chemical is required to produce each of the harmful effects, how we recognize an overexposure, and how to treat it.
One way to see if a chemical will hurt people is to learn how the chemical is absorbed, used, and released by the body; for some chemicals, animal testing may be necessary. Animal testing may also be used to identify health effects such as kidney or liver damage, cancer, or birth defects. Without laboratory animals, scientists would lose a basic method to get information needed to make wise decisions to protect public health. Scientists have the responsibility to treat research animals with care and compassion. Laws today protect the welfare of research animals, and scientists must comply with strict animal care guidelines.
Uranium is a chemical substance that is also radioactive. Scientists have never detected harmful radiation effects from low levels of natural uranium, although some may be possible. However, scientists have seen chemical effects. A few people have developed signs of kidney disease after intake of large amounts of uranium. Animals have also developed kidney disease after they have been treated with large amounts of uranium, so it is possible that intake of a large amount of uranium might damage your kidneys. There is also a chance of getting cancer from any radioactive material like uranium. Natural and depleted uranium are only weakly radioactive and are not likely to cause you to get cancer from their radiation. No human cancer of any type has ever been seen as a result of exposure to natural or depleted uranium. Uranium can decay into other radionuclides, which can cause cancer if you are exposed to enough of them for a long enough period. Doctors that studied lung and other cancers in uranium miners did not think that uranium radiation caused these cancers. The miners smoked cigarettes and were exposed to other substances that we know cause cancer, and the observed lung cancers were attributed to large exposures to radon and its radioactive transformation products.
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The chance of getting cancer is greater if you are exposed to enriched uranium, because it is more radioactive than natural uranium. Cancer may not become apparent until many years after a person is exposed to a radioactive material (from swallowing or breathing it). Just being near uranium is not dangerous to your health because uranium gives off very little of the penetrating gamma radiation. However, uranium is normally accompanied by the other transformation products in its decay chain, so you would be exposed to their radiation as well.
The Committee on the Biological Effects of Ionizing Radiation (BEIR IV) reported that eating food or drinking water that has normal amounts of uranium will most likely not cause cancer or other health problems in most people. The Committee used data from animal studies to estimate that a small number of people who steadily eat food or drink water that has larger-than-normal quantities of uranium in it could get a kind of bone cancer called a sarcoma. The Committee reported calculations showing that if people steadily eat food or drink water containing about 1 pCi of uranium every day of their lives, bone sarcomas would be expected to occur in about 1 to 2 of every million people after 70 years, based on the radiation dose alone. However, we do not know this for certain because people normally ingest only slightly more than this amount each day, and people who have been exposed to larger amounts have not been found to get cancer.
We do not know if exposure to uranium causes reproductive effects in people. Very high doses of uranium have caused reproductive problems (reduced sperm counts) in some experiments with laboratory animals. Most studies show no effects. For more information about how uranium can affect your health, see Chapter 2.
1.6 HOW CAN URANIUM AFFECT CHILDREN?
This section discusses potential health effects from exposures during the period from conception to maturity at 18 years of age in humans. Potential effects on children resulting from exposures of the parents are also considered.
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Like adults, children are exposed to small amounts of uranium in air, food, and drinking water. However, no cases have been reported where exposure to uranium is known to have caused health effects in children. It is possible that if children were exposed to very high amounts of uranium they might have damage to their kidneys like that seen in adults. We do not know whether children differ from adults in their susceptibility to health effects from uranium exposure.
It is not known if exposure to uranium has effects on the development of the human fetus. Very high doses of uranium in drinking water can affect the development of the fetus in laboratory animals. One study reported birth defects and another reported an increase in fetal deaths. However, we do not believe that uranium can cause these problems in pregnant women who take in normal amounts of uranium from food and water, or who breathe the air around a hazardous waste site that contains uranium.
Very young animals absorb more uranium into their blood than adults when they are fed uranium. We do not know if this happens in children.
Measurements of uranium have not been made in pregnant women, so we do not know if uranium can cross the placenta and enter the fetus. In an experiment with pregnant animals, only a very small amount (0.03%) of the injected uranium reached the fetus. Even less uranium is likely to reach the fetus in mothers exposed by inhaling, swallowing, or touching uranium. No measurements have been made of uranium in breast milk. Because of the chemical properties of uranium, it is unlikely that it would concentrate in breast milk.
1.7 HOW CAN FAMILIES REDUCE THE RISK OF EXPOSURE TO URANIUM?
If your doctor finds that you have been exposed to significant amounts of uranium, ask whether your children might also be exposed. Your doctor might need to ask your state health department to investigate.
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It is possible that higher-than-normal levels of uranium may be in the soil at a hazardous waste site. Some children eat a lot of dirt. You should prevent your children from eating dirt. Make sure they wash their hands frequently, and before eating. If you live near hazardous waste site, discourage your children from putting their hands in their mouths or from engaging in other hand-to-mouth activities.
1.8 IS THERE A MEDICAL TEST TO DETERMINE WHETHER I HAVE BEEN EXPOSED TO URANIUM?
Yes, there are medical tests that can determine whether you have been exposed by measuring the amount of uranium in your urine, blood, and hair. Urine analysis is the standard test. If you take into your body a larger-than-normal amount of uranium over a short period, the amount of uranium in your urine may be increased for a short time. Because most uranium leaves the body within a few days, normally the amount in the urine only shows whether you have been exposed to a larger-than-normal amount within the last week or so. If the intake is large or higher-thannormal levels are taken in over a long period, the urine levels may be high for a longer period of time. Many factors can affect the detection of uranium after exposure. These factors include the type of uranium you were exposed to, the amount you took into your body, and the sensitivity of the detection method. Also, the amount in your urine does not always accurately show how much uranium you have been exposed to. If you think you have been exposed to elevated levels of uranium and want to have your urine tested, you should do so promptly while the levels may still be high. In addition to uranium, the urine could be tested for evidence of kidney damage, by looking for protein, glucose, and nonprotein nitrogen, which are some of the chemicals that can appear in your urine because of kidney damage. Testing for these chemicals could determine whether you have kidney damage. However, since kidney damage is also caused by several common diseases, such as diabetes, it would not tell you if the damage was caused by the presence of uranium in your body.
A radioactivity counter can tell if your skin is contaminated with uranium, because uranium is radioactive. If you inhale large amounts of uranium, it may be possible to measure the amount
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of radioactivity in your body with special radiation measurement instruments. See Chapters 2 and 6 for more information.

1.9 WHAT RECOMMENDATIONS HAS THE FEDERAL GOVERNMENT MADE TO PROTECT HUMAN HEALTH?
International and national organizations like the International Commission on Radiological Protection (ICRP) and the National Council on Radiation Protection and Measurements (NCRP) provide recommendations for protecting people from materials, like uranium, that give off ionizing radiation. The federal government considers these recommendations and develops regulations and guidelines to protect public health. Regulations can be enforced by law. Federal agencies that develop regulations for toxic substances include the EPA, the Nuclear Regulatory Commission (NRC), the Occupational Safety and Health Administration (OSHA), and the Food and Drug Administration (FDA). Recommendations provide valuable guidelines to protect public health but cannot be enforced by law. Federal organizations that develop recommendations for toxic substances include the Agency for Toxic Substances and Disease Registry (ATSDR) and the National Institute for Occupational Safety and Health (NIOSH).
Regulations and recommendations can be expressed as levels that are not to be exceeded in air, water, soil, or food that are usually based on levels that affect animals. Then they are adjusted with appropriate safety factors to help protect people. Sometimes these not-to-exceed levels differ among federal organizations because of different exposure times (an 8-hour workday or a 24-hour day), the use of different animal studies, or other factors.
Recommendations and regulations are also periodically updated as more information becomes available. For the most current information, check with the federal agency or organization that provides it. Some regulations and recommendations for uranium are discussed below.
EPA has not set a limit for uranium in air, but it has set a goal of no uranium in drinking water. EPA calls this the Maximum Contaminant Level Goal (MCLG), but recognizes that, currently, there is no practical way to meet this goal. Because of this, EPA proposed in 1991 to allow up to
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20 μg of uranium per liter (20 μg/L) in drinking water, and states began regulating to achieve this level. EPA calls this the Maximum Contaminant Level (MCL). The MCL for uranium is based on calculations that if 150,000 people drink water that contains 20 μg/L of uranium for a lifetime, there is a chance that one of them may develop cancer from the uranium in the drinking water. In 1994, EPA considered changing the MCL to 80 μg per liter based on newer human intake and uptake values and the high cost of reducing uranium levels in drinking water supplies. In 1998, EPA temporarily dropped its 1991 limit, but is currently working to develop an appropriate limit based on a broader range of human and animal health studies. ATSDR, other federal agencies, Canada, and other professionals are advising EPA regarding a new MCL. Canada is currently developing its own national guideline value because that country has the richest known uranium ore deposits in the world and high uranium concentrations in some of its well water.
EPA has also decided that any accidental uranium waste containing 0.1 curies of radioactivity (150 kilograms) must be cleaned up. EPA calls this the Reportable Quantity Accidental Release. EPA also has established a standard for uranium mill tailings. In the workplace, NIOSH/OSHA has set a Recommended Exposure Limit (REL) and a Permissible Exposure Limit (PEL) of
1 mg/m3 (34 pCi/m3) for uranium dust, while the NRC has an occupational limit of 0.2 mg/m3 (130 pCi/m3). The NRC has set uranium release limits at 0.06 pCi/m3 (0.09 μg/m3) of air and 300 pCi/liter (450 μg/liter) of water. NRC and OSHA expect that the public will normally be exposed to much lower concentrations. For more information about recommendations the federal government has made to protect your health, see Chapter 7.
2 WHERE CAN I GET MORE INFORMATION?

If you have any more questions or concerns, please contact your community or state health or environmental quality department or
Agency for Toxic Substances and Disease RegistryDivision of Toxicology1600 Clifton Road NE, Mailstop E-29Atlanta, GA 30333
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* Information line and technical assistance
Phone: 1-888-42-ATSDR (1-888-422-8737)Fax: (404) 639-6315 or 6324
ATSDR can also tell you the location of occupational and environmental health clinics. These clinics specialize in recognizing, evaluating, and treating illnesses resulting from exposure to hazardous substances.
* To order toxicological profiles, contact
National Technical Information Service5285 Port Royal RoadSpringfield, VA 22161Phone: (800) 553-6847 or (703) 605-6000
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2.1 INTRODUCTION
The primary purpose of this chapter is to provide public health officials, physicians, toxicologists, and other interested individuals and groups with an overall perspective of the toxicology of uranium. It contains descriptions and evaluations of toxicological studies and epidemiologic investigations and provides conclusions, where possible, on the relevance of toxicity and toxicokinetic data to public health.
A glossary and list of acronyms, abbreviations, and symbols can be found at the end of this profile.
The health effects associated with oral or dermal exposure to natural and depleted uranium appear to be solely chemical in nature and not radiological, while those from inhalation exposure may also include a slight radiological component, especially if the exposure is protracted. A comprehensive review by the Committee on the Biological Effects of Ionizing Radiation (BEIR IV) concluded that ingesting food or water containing normal uranium concentrations will most likely not be carcinogenic or cause other health problems in most people. Inhaled uranium is associated with only a low cancer risk, with the main risk being associated with the co-inhalation of other toxic and/or carcinogenic agents, such as the radioactive transformation products of radon gas and cigarette smoke. Very high oral doses of uranium have caused renal damage in humans. Animal studies in a number of species and using a variety of compounds confirm that uranium is a nephrotoxin and that the most sensitive organ is the kidney. Hepatic and developmental effects have also been noted in some animal studies. This profile is primarily concerned with the effects of exposure to natural and depleted uranium, but does include limited discussion regarding enriched uranium, which is considered to be more of a radiological than a chemical hazard. Also, whenever the term “radiation” is used, it applies to ionizing radiation and not to nonionizing radiation.
Although natural and depleted uranium are primarily chemical hazards, the next several paragraphs describe the radiological nature of the toxicologically-important uranium isotopes, because individual isotopes are addressed in some of the health effects studies. Uranium is a naturally occurring radioactive element and a member of the actinide series. Radioactive elements are those that undergo spontaneous transformation (decay), in which energy is released (emitted) either in the form of particles, such as alpha or beta particles, or electromagnetic radiation with energies sufficient to cause ionization, such as gamma or X-rays. This transformation or decay results in the formation of different elements, some of which
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may themselves be radioactive, in which case they will also decay. The process continues until a stable (nonradioactive) state is reached (see Appendix D for more information).
Uranium exists in several isotopic forms, all of which are radioactive. The most toxicologically important of the 22 currently recognized uranium isotopes are anthropogenic uranium-232 (232U) and uranium-233 (233U) and naturally occurring uranium-234 (234U), uranium-235 (235U), and uranium-238 (238U). When an atom of any of these five isotopes decays, it emits an alpha particle (the nucleus of a helium atom) and transforms into a radioactive isotope of another element. The process continues through a series of radionuclides until reaching a stable, non-radioactive isotope of lead. The radionuclides in these transformation series (such as radium and radon), emit alpha, beta, and gamma radiations with energies and intensities that are unique to the individual radionuclide.
Natural uranium consists of isotopic mixtures of 234U, 235U, and 238U. There are three kinds of mixtures (based on the percentage of the composition of the three isotopes): natural uranium, enriched uranium, and depleted uranium. Natural uranium, including uranium ore, is comprised of 99.284% 238U, 0.711% 235U, and 0.005% 234U by mass. Combining these mass percentages with the unique half-life of each isotope converts mass into radioactivity units and shows that uranium ore contains 48.9% 234U, 2.25% 235U, and 48.9% 238U by radioactivity, and has a very low specific activity of 0.68 μCi/g (Parrington et al. 1996). Enriched and depleted uranium are the products of a process which increases (or enriches) the percentages of 234U and 235U in one portion of a uranium sample and decreases (or depletes) their percentages in the remaining portion. Enriched uranium is quantified by its 235U percentage Uranium enrichment for nuclear energy produces uranium that typically contains 3% 235U. Uranium enrichment for a number of other purposes, including nuclear weapons, can produce uranium that contains as much as 97.3% 235U and has a higher specific activity (.50 μCi/g). The residual uranium after the enrichment process is called “depleted uranium” (DU), which possesses even less radioactivity (0.36 μCi/g) than natural uranium. The Nuclear Regulatory Commission (NRC) considers the specific activity of depleted uranium to be 0.36 μCi/g (10 CFR 20), but more aggressive enrichment processes can drive this value slightly lower (0.33 μCi/g). In this profile, both natural and depleted uranium are referred to as “uranium.” The higher specific-activity mixtures and isotopes are described in the profile as “enriched uranium,” or as 232U, 233U, or 234U, as applicable, in the summary of the studies in which these mixtures and isotopes were used.
Because uranium is a predominantly alpha-emitting radionuclide, there is a concern for potential DNA damage and fragmentation if alpha particles reach cell nuclei. Attempts by cells to repair this
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fragmentation, if it occurs, may result in repair errors, producing gene mutations or chromosomal aberrations. These effects, when sufficiently severe, may be manifested as cancer and possibly as developmental malformations. However, the genetic effects of radiation have not been observed in humans with exposure to radiation, including that from uranium. Ionizing radiation may also promote carcinogenesis by an apoptotic mechanism by which radiation-induced cell death in tissues or organs elicits an increased cell proliferation response to replace the lost cells. Increased mitotic activity may afford cancer cells a preferential advantage in clonal expansion.
Although radiation exposure has been generally assumed to be carcinogenic at all dose levels, no correlation has been established at low doses such as occur from exposure to natural radiation background levels. This is largely attributable to two factors: (1) it is difficult to construct and obtain meaningful data from epidemiological studies where exposure is near background exposure levels, and (2) the data are not statistically significant enough to substantiate a detectable health impact. Recent risk assessment reviews of carcinogenicity and exposure to hazardous chemicals, including radiation, have been questioning the non-threshold assumption. With specific reference to radiation, there is increasing biological evidence that there is a threshold for radiation-induced carcinogenicity (Clark 1999).
The National Research Council Committee on the Biological Effects of Ionizing Radiation BEIR IV report stated that ingesting uranium in food and water at the naturally occurring levels will not cause cancer or other health problems in people. However, based on the zero-threshold linear dose-response model (a conservative model that is inherently unverifiable and is intended to be used as an aid to risk-benefit analysis and not for predicting cancer deaths), the BEIR IV committee calculated that the ingestion of an additional 1 pCi/day (0.0015 mg/day) of soluble natural uranium would lead to a fractional increase in the incidence rate of osteogenic sarcoma (bone cancer) of 0.0019. This means that over a period of 70 years (the nominal lifetime length), if everyone were exposed at that level, the number of bone cancer cases in a U.S. population of 250 million would increase from 183,750 to about 184,100. Currently, there are no unequivocal studies that show that intake of natural or depleted uranium can induce radiation effects in humans or animals. The available information on humans and animals suggests that intake of uranium at the low concentrations usually ingested by humans or at levels found at or near hazardous waste sites is not likely to cause cancer. The BEIR IV committee, therefore, concluded that “…exposure to natural uranium is unlikely to be a significant health risk in the population and may well have no measurable effect.”
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Exposure to enriched uranium, used as uranium fuel in nuclear energy production, may present a radiological health hazard. Although uranium-associated cancers have not been identified in humans, even following exposure to highly enriched uranium, higher doses associated with highly enriched, high specific activity uranium may be able to produce bone sarcomas in humans. Evidence from animal studies suggests that high radiation doses associated with large intakes of 234U and 235U-enriched uranium compounds can be hazardous. Adverse effects reported from such exposures include damage to the interstitium of the lungs (fibrosis) and cardiovascular abnormalities (friable vessels). However, access to 235U-enriched or other high specific-activity uranium is strictly regulated by the NRC and the U.S. Department of Energy (DOE). Therefore, the potential for human exposure to this level of radioactivity is limited to rare accidental releases in the workplace.
The potential for adverse noncancerous radiological health effects from uranium is dependent on several factors, including physicochemical form and solubility, route of entry, distribution in the various body organs, the biological retention time in the various tissues, and the energy and intensity of the radiation. The potential for such effects is generally thought to be independent of the known chemical toxicity of uranium. While the chemical properties affect the distribution and biological half-life of a radionuclide, the damage from radiation is independent of the source of that radiation. In this profile, there is little, or equivocal, specific information regarding the influence of radiation from uranium on certain biological effect end points in humans, such as reproductive, developmental, or carcinogenic effects. There is evidence, however, from the large body of literature concerning radioactive substances that alpha radiation can affect these processes in humans (see Appendix D for additional information on the biological effects of radiation). However, because the specific activities of natural and depleted uranium are low, no radiological health hazard is expected from exposure to natural and depleted uranium. Since the radiological component of natural uranium has essentially been discounted as a significant source of health effects, this leaves only the chemical effects of uranium to contend with. The chemical (nonradiological) properties of natural uranium and depleted uranium are identical; therefore, the health effects exerted by each are expected to be the same. The results of the available studies in humans and animals are consistent with this conclusion. The potential health impacts of depleted uranium are specifically addressed in a recent Department of Energy publication (DOE 1999).
Uranium is a heavy metal that forms compounds and complexes of different varieties and solubilities. The chemical action of all isotopes and isotopic mixtures of uranium is identical, regardless of the specific activity (i.e., enrichment), because chemical action depends only on chemical properties. Thus, the chemical toxicity of a given amount or weight of natural, depleted, and enriched uranium is identical.
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The toxicity of uranium varies according to its chemical form and route of exposure. On the basis of the toxicity of different uranium compounds in animals, it was concluded that the relatively more water-soluble compounds (uranyl nitrate hexahydrate, uranium hexafluoride, uranyl fluoride, uranium tetrachloride, uranium pentachloride) were the most potent renal toxicants. The less water-soluble compounds (sodium diuranate, ammonium diuranate) were of moderate-to-low renal toxicity, and the insoluble compounds (uranium tetrafluoride, uranium trioxide, uranium dioxide, uranium peroxide, triuranium octaoxide) had little potential to cause renal toxicity but could cause pulmonary toxicity when exposure was by inhalation.
The terms soluble, moderately soluble, and insoluble are often used in this profile without relisting the specific compounds. Generally, hexavalent uranium, which tends to form soluble compounds, is more likely to be a systemic toxicant than tetravalent uranium, which forms insoluble compounds. Ingested uranium is less toxic than inhaled uranium, which may be partly attributable to the relatively low gastrointestinal absorption of uranium compounds. Only <0.1–6% of even the more soluble uranium compounds are absorbed in the gastrointestinal tract (Leggett 1989). The available data on a variety of uranium compounds are sufficient to conclude that uranium has a low order of metallotoxicity (chemical toxicity) in humans. This low order results from the high exposures to which animals in these studies were exposed without adverse effects in many cases. The ICRP (1995) recommends a gastrointestinal absorption fraction of 0.02 (i.e. 2%) for uranium ingested in relatively stable form.
The hazard from inhaled uranium aerosols, or any noxious agent, is determined by the likelihood that the agent will reach the site of its toxic action. Two main factors that influence the degree of hazard from toxic airborne particles are the site of deposition in the respiratory tract of the particles and the fate of the particles within the lungs. The deposition site within the lungs depends mainly on the particle size of the inhaled aerosol, while the subsequent fate of the particle depends mainly on the physical and chemical properties of the inhaled particles and the physiological status of the lungs.
Human and animal studies have shown that long-term retention in the lungs of large quantities of inhaled insoluble uranium particles (e.g., carnotite dust [4% uranium as uranium dioxide and triuranium octaoxide, 80–90% quartz, and rat > guinea pig > pig > mouse (Orcutt 1949).
2.2.1.1 Death
The lethal effects of inhalation exposure to uranium have been investigated in humans in epidemiological studies and in animal studies under controlled conditions. Epidemiological studies indicate that routine exposure of humans (in the workplace and the environment at large) to airborne uranium is not associated with increased mortality. Brief accidental exposures to very high concentrations of uranium hexafluoride have caused fatalities in humans. Laboratory studies in animals indicate that inhalation exposure to certain uranium compounds can be fatal. These deaths are believed to result from renal failure caused by absorbed uranium. The low specific activity of uranium precludes the possibility of absorbing enough uranium to deliver a lethal dose of radiation.
No definitive evidence has been found in epidemiologic studies that links human deaths to uranium exposure. Among uranium miners, death rates from diseases of the cardiovascular system and the urogenital system were decreased compared to other populations. Uranium miners have higher-thanexpected rates of death from lung cancer; however, this finding is attributed to the radiological effects of radon and its decay products, which are progeny of uranium and, therefore, present in uranium mines. In addition, the role of tobacco smoking in these deaths was not evaluated (Archer et al. 1973a; Gottlieb and Husen 1982; Lundin et al. 1969; Samet et al. 1984, 1986). Epidemiologic studies of workers at uranium mill and metal processing plants (where there is little or no exposure to radon in excess of normal environmental levels) showed no increase in overall deaths attributable to exposure to uranium (Archer et al. 1973b; Brown and Bloom 1987; Checkoway et al. 1988; Cragle et al. 1988; Hadjimichael et al. 1983; Polednak and Frome 1981; Scott et al. 1972; Waxweiler et al. 1983).
Deaths occurred after accidental releases of uranium hexafluoride at uranium-processing facilities in 1944 and 1986, but these deaths were not attributed to the uranium component of this compound (Kathren and
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Moore 1986; Moore and Kathren 1985; USNRC 1986). These releases resulted in the generation of concentrated aerosols of highly toxic hydrofluoric acid and uranyl fluoride2. In the 1944 incident exposure time was estimated to be only 17 seconds, deaths occurred in 2 of 20 workers within an hour and were attributed to severe chemical burns of the lungs. In the 1986 incident, 1 of 23 workers died from massive pulmonary edema, indicating that inhalation of hydrofluoric acid was responsible for death. Estimated airborne concentrations were 20 mg uranium hexafluoride/m3 for a 1-minute exposure and 120 mg uranium hexafluoride/m3 for a 60-minute exposure (15.2 and 91 mg U/m3, respectively).
Mortality can be induced in animals exposed to sufficiently high concentrations of pure uranium compounds. The acute-duration LC50 (lethal concentration, 50% death) for uranium hexafluoride has been calculated for Long-Evans rats and Hartley guinea pigs (Leach et al. 1984). The animals were exposed to uranium hexafluoride in a nose-only exposure apparatus for periods of up to 10 minutes and then observed for 14 days. Total mortality in rats was 34% (157/460): 25% of the deaths occurred during exposure or within 48 hours, 59% between days 3 and 7, and 17% between days 7 and 14. Guinea pigs were more sensitive than rats; total mortality was 46% (36/78), and 64% of deaths occurred within 48 hours. In guinea pigs, the LC50 was estimated as 35,011 mg U/m3 for a 2-minute exposure limit. For a 5-minute inhalation exposure, the LC50 in Long-Evans rats was estimated as 26,098 mg U/m3; the LC50 for a 10-minute inhalation was estimated as 8,114 mg U/m3. The animals that died showed some damage to the respiratory tract, probably due to hydrofluoric acid, but this damage was not judged to be the cause of death, at least in the animals that died more than 2 days postexposure. Urinalysis and histopathological examination indicated that renal injury was the primary cause of death (Leach et al. 1984). In other acute lethality studies, rats, mice, and guinea pigs suffered 10, 20, and 13% mortality, respectively, following a 10-minute inhalation of uranium hexafluoride corresponding to 637 mg U/m3 (Spiegl 1949).
In intermediate-duration studies, rabbits and cats were generally the most sensitive species to uranium lethality. Deaths in these studies generally occurred beginning 2 weeks after exposure started and continued to the end of the experiment. Exposure to 2 mg U/m3 (as uranium hexafluoride) 6 hours a day for 30 days caused 5, 20, and 80% mortality in guinea pigs, dogs, and rabbits, respectively (Spiegl 1949). An exposure to 9.5 mg U/m3 (as uranyl nitrate hexahydrate) for 8 hours per day, 5 days per week for 30 days caused 10% mortality in rats and guinea pigs, and 75% mortality in dogs. Exposure to 2 mg U/m3 killed all four cats tested (Roberts 1949). Exposure to 9.2 mg U/m3 (as uranyl fluoride) 6 hours a day, 5.5 days a week for 5 weeks caused 0%, 100%, 83%, and 55% mortality in rats, mice, rabbits and guinea pigs, and deaths in two dogs and two cats tested at this concentration (Rothstein
2 Uranium hexafluoride rapidly dissociates into hydrofluoric acid and uranyl fluoride on contact with moisture in the air.
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1949a). The lowest exposure causing death with uranyl fluoride was 0.15 mg U/m3 in mice and rabbits and 2.2 mg U/m3 in guinea pigs. Exposure to 15.4 mg U/m3 (as uranium peroxide) 5 hours a day, 5 days a week for 23 days caused 10, 63, 40, 80, and 100% mortality in rats, mice, guinea pigs, rabbits, and cats, respectively, while 9.2 mg U/m3 killed all the dogs tested (Dygert 1949d). Inhalation of air containing 15 mg U/m3 (as sodium diuranate) for 6 hours a day, 5.5 days a week for 5 weeks caused 13 and 28% mortality in guinea pigs and rabbits, respectively (Rothstein 1949d).
Insoluble uranium compounds were also lethal to animals by the inhalation route, but at higher concentrations than with soluble compounds. Exposure to 15.8 mg U/m3 (as uranium trioxide) 6 hours a day, 5.5 days a week for 4 weeks caused 10, 9, 17, and 67% mortality in rats, guinea pigs, dogs, and rabbits, respectively (Rothstein 1949c). Inhalation of air containing 19.4 mg U/m3 (as uranium dioxide) for 6 hours a day, 5.5 days a week for 5 weeks, caused 60% mortality in rabbits but no mortality in rats, mice, guinea pigs, or dogs (Rothstein 1949b). Inhalation of air containing 18 mg U/m3 (as uranium tetrafluoride) for 5 hours a day for 30 days caused 15, 32, 33, and 100% mortality in guinea pigs, rats, rabbits, and cats, respectively, and death in a single dog tested at this concentration. Inhalation at 4 mg U/m3 caused no deaths in a group of 18 dogs, and one death in a group of 30 rats (Dygert 1949a). A mortality of 4% was observed among rabbits given 3 mg U/m3 (Stokinger et al. 1953). Exposure to
6.8 mg U/m3 (as ammonium diuranate) 6 hours a day for 30 days caused 20 and 100% mortality in guinea pigs and rabbits, respectively (Dygert 1949b).
In chronic-duration experiments, inhalation of 2 mg U/m3 as uranyl nitrate hexahydrate for 6 hours a day,
5.5 days a week for 92–100 weeks resulted in 1% mortality in rats (Stokinger et al. 1953). This is not an unusual mortality rate for rats, so it is unlikely that these deaths can be attributed to uranium exposure. Dogs exposed to uranyl nitrate hexahydrate for 2 years suffered 4% mortality (Stokinger et al. 1953). One dog died at 0.25 mg U/m3, and another at 2 mg U/m3 out of 25 exposed dogs. Death may or may not have been attributable to uranium, according to the study investigator.
In several other inhalation-exposure animal studies, no deaths were observed when either soluble or insoluble uranium compounds were administered. In one of these animal studies, no mortality was observed in monkeys exposed by inhalation to uranium dioxide dust at a concentration of 5 mg U/m3 for 5 years. The death of Beagle dogs similarly exposed could not be attributed to uranium by the investigators (Leach et al. 1970).
The percent mortality values for each species and other LOAEL values for mortality from exposure to uranium by the inhalation route are presented in Table 2-1 and plotted in Figure 2-1.
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2.2.1.2 Systemic Effects
No human studies were located regarding the cardiovascular, musculoskeletal, endocrine, metabolic, dermal, ocular, body weight, or other systemic effects of elemental uranium following acute-duration inhalation exposure. Nor were any human studies located regarding the respiratory, hematological, cardiovascular, gastrointestinal, musculoskeletal, hepatic, renal, endocrine, metabolic, dermal, ocular, body weight, or other systemic effects of uranium following intermediate-duration inhalation exposure. No studies were found regarding the cardiovascular, gastrointestinal, musculoskeletal, renal, endocrine, metabolic, dermal, ocular, body weight, or other systemic effects in humans following chronic-duration inhalation exposure. The existing human data on the respiratory and hepatic effects of uranium are limited to acute- and chronic-duration inhalation exposures, hematological effects are limited to chronic-duration inhalation exposure, and gastrointestinal and renal effects are limited to acute-duration inhalation exposure.
No animal studies were located regarding the endocrine, metabolic, dermal, or ocular effects of uranium in animals following acute-duration inhalation exposures to uranium. Nor were any studies located regarding the metabolic, dermal, ocular, or other systemic effects in animals following intermediate-duration inhalation exposure to uranium. There are animal data for acute-, intermediate-, and chronic-duration inhalation exposures to uranium for respiratory, hematological, cardiovascular, gastrointestinal, renal, or body weight effects. However, animal data on hepatic effects are limited to acute- and chronic-duration inhalation exposures to uranium.
The highest NOAEL values and all reliable LOAEL values in each species and duration category for systemic effects from chemical exposure to uranium by the inhalation route are presented in Table 2-1 and plotted in Figure 2-1. The radiation effect level values in each species and duration category for systemic effects from radiation exposure to uranium by the inhalation route are presented in Table 2-2 and plotted in Figure 2-2.
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Respiratory Effects. The hazard from inhaled uranium aerosols, or from any noxious agent, is the likelihood that the agent will reach the site of its toxic action. Two of the main factors that influence the degree of hazard from toxic airborne particles are 1) the site of deposition in the respiratory tract of the particles and 2) the fate of the particles within the lungs. The deposition site within the lungs depends mainly on the particle size of the inhaled aerosol, while the subsequent fate of the particle depends mainly on the physical and chemical properties of the inhaled particles and the physiological status of the lungs.
Small particles (about 2 micrometers [μm] or smaller in diameter) tend to be deposited in the alveoli. The alveoli, frequently called the “deep respiratory tract,” form the functional part of the lungs where gas exchange occurs. As the particle size increases, progressively fewer particles penetrate into the deep respiratory tract, and increasingly greater fractions of the inhaled particles are deposited in the upper respiratory tract. The respiratory tract is a system of ducts that starts at the nares and includes the pharynx, larynx, trachea, and a complex series of bronchi and bronchioles that terminate in several thousand alveoli. Three different mechanisms are involved in the removal of particles from the respiratory tract. The first is mucociliary action in the upper respiratory tract (trachea, bronchi, bronchioles, and terminal bronchioles), which sweeps particles deposited there into the throat, where they are either swallowed into the gastrointestinal tract or spat out. The two other clearance mechanisms, dissolution (which leads to absorption into the bloodstream) and phagocytosis (removal by specialized cells in the process), deal mainly with the particles deposited in the deep respiratory tract (respiratory bronchioles, alveolar ducts, and alveolar sacs) (ICRP 1994; NCRP 1997). The less soluble uranium particles may remain in the lungs and in the regional lymph nodes for weeks (uranium trioxide, uranium tetrafluoride, uranium tetrachloride) to years (uranium dioxide, triuranium octaoxide).
In acute exposures, respiratory disease may be limited to interstitial inflammation of the alveolar epithelium, leading eventually to emphysema or pulmonary fibrosis (Cooper et al. 1982; Dungworth 1989; Stokinger 1981; Wedeen 1992). In studies of the pulmonary effects of airborne uranium dust in uranium miners and in animals, the respiratory diseases reported are probably aggravated by the inhalable dust particles’ (the form in which uranium is inhaled) toxicity because most of the respiratory diseases reported in these studies are consistent with the effects of inhaled dust (Dockery et al. 1993). In some of these instances, additional data from the studies show that the workers were exposed to even more potent respiratory tract irritants, such as silica and vanadium pentaoxide (Waxweiler et al. 1983).
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The effects of massive acute exposures to uranium in humans, as well as epidemiologic or clinical studies of uranium mine workers chronically exposed to mine atmospheres (containing other noxious agents that include silica, diesel fumes, cigarette smoke, and radon and its daughters), have been investigated. Several epidemiologic studies have reported respiratory diseases in uranium mine and mill workers, who are also exposed to significant amounts of dust and other pulmonary irritants, but not in uranium-processing workers, who are not exposed to these potential aggravants.
Accidental exposure of workers to estimated airborne concentrations of 20 mg uranium hexafluoride/m3 for a 1-minute exposure and 120 mg uranium hexafluoride/m3 for a 60-minute exposure (15.2 and 91 mg U/m3, respectively) resulted in acute respiratory irritation, which is attributed to the hydrofluoric acid decomposition product. One worker died of pulmonary edema a few hours after the accident (Fisher et al. 1990; USNRC 1986). In another report, 20 men who were seriously injured following accidental exposure to a stream of uranium hexafluoride when a transportation cask ruptured showed signs of pulmonary edema, which also was attributed to hydrofluoric acid. After 3 weeks, most had normal clinical findings and were considered to be in excellent health. A follow-up examination 38 years later on three of the injured workers showed no detectable uranium deposition and no respiratory findings attributable to the exposure (Kathren and Moore 1986). No clinical signs of pulmonary toxicity were found in about 100 uranium-processing workers exposed to insoluble uranium dust at levels of 0.5–2.5 mg U/m3 for about 5 years (Eisenbud and Quigley 1955). Other reports of workers in the uranium processing industry did not show increased deaths due to diseases of the respiratory system related to exposure to uranium (Brown and Bloom 1987; Cragle et al. 1988; Polednak and Frome 1981; Scott et al. 1972).
A 30-year follow-up study in which ionizing radiation hazard was assessed for a study cohort consisting of 995 workers in a uranium-processing facility that operated between 1943 and 1949 found statistically significant increases in death from all causes. Significantly increased mortality was observed for cancer of the larynx and for pneumonia, but not for lung cancer. The workers were exposed to internal radiation from the inhaled uranium dust, with an upper limit of 1,000 mSv. The data (external radiation badge) for the last 24 months of operation indicated that the highest cumulative external gamma dose for a worker was about 20 mSv. Long-term occupational exposure was evaluated in a subcohort that received 150 mSv/year or more. Because the workers were also exposed to radon-222 (222Ra), chlorine, hydrofluoric acid, lead sulfate, nickel, nitric acid and nitrogen oxides, silicon dioxide, and sulfuric acid,
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the etiology of the reported laryngeal disease is uncertain (Dupree et al. 1987). An increased incidence of deaths (Standard Mortality Ratio [SMR] =2.29) from obstructive pulmonary disease was found in 4,106 workers in a nuclear fuels fabrication plant who were employed for more than 6 months from 1956 to 1978 (Hadjimichael et al. 1983). However, the overall death rate and rate of all cancers combined were lower than expected. The association of disease with exposure to uranium was not confirmed.
The pulmonary toxicity of uranium compounds varies in animals. Reports of pulmonary toxicity in animals after acute-duration exposure to uranium are limited to experiments with uranium hexafluoride. Gasping and severe irritation to the nasal passages were reported after 10 minute exposures at 637 mg U/mg3 in rats and mice (Spiegl 1949) and nasal hemorrhage in rats after a 5 minute exposure to 54,503 mg/m3 (Leach et al. 1984). Uranium hexafluoride promptly hydrolyzes on contact with water to uranyl fluoride and hydrofluoric acid. Thus, the animals were potentially exposed to hydrofluoric acid, a potent toxicant to respiratory tract epithelium, which probably contributed to pulmonary tissue destruction (Leach et al. 1984; Spiegl 1949; Stokinger et al. 1953). In addition, exposure to fluoride ions can result in hypocalcemia, hypomagnesemia, pulmonary edema, metabolic acidosis, ventricular arrhythmia, and death (Meditext 1998).
Intermediate-duration exposure to uranium compounds also caused pulmonary toxicity, particularly when exposure was to uranium hexafluoride. Exposure of rats, mice, and guinea pigs to this compound for 6 hours/day for 30 days at 13 mg U/m3 resulted in pulmonary edema, hemorrhage, emphysema, and inflammation of the bronchi and alveoli (Spiegl 1949). Milder effects were observed with other uranium compounds in a series of experiments where exposure conditions were similar to those found in the workplace (i.e., 5–6 hours/day, 5–6 days/week). For example, rhinitis was observed in cats and dogs after 30 days exposure to 18 mg U/m3 as uranium tetrafluoride (Dygert 1949a) and after 5 weeks exposure to
9.2 mg U/m3 as uranyl fluoride (Rothstein 1949a). Histopathological evidence of toxicity was observed in several studies, including slight degenerative changes in rats and dogs exposed to 16 mg U/m3 as uranium trioxide (Rothstein 1949c) and dogs exposed to 9.5 mg U/m3 as uranyl nitrate (Roberts 1949). Uranium dioxide and triuranium octaoxide did not cause toxicity (Dygert 1949c; Rothstein 1949b). Carnotite uranium ore did not cause toxicity in mice or guinea pigs, but hemorrhagic lungs were observed in dogs (Pozzani 1949). The species differences may reflect deeper penetration of this material into the dog respiratory tract. Rabbits were more sensitive to respiratory effects of uranium compounds than other species. Severe respiratory effects (pulmonary edema, hemorrhage) were observed in this species with
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exposure to 6.8 mg U/m3 as ammonium diuranate (Dygert 1949b), 15.4 mg U/m3 as uranium peroxide (Dygert 1949d), 16 mg U/m3 as uranium trioxide (Rothstein 1949c) and 22 mg U/m3 as carnotite uranium ore (Pozzani 1949). Uranium dioxide at 19.4 mg U/m3 did not cause respiratory effects in rabbits (Rothstein 1949b).
In chronic-duration exposure tests, a total of 3,100 test animals, including rats, rabbits, guinea pigs, and dogs were exposed to aerosols containing 0.05–10 mg U/m3 of various uranium compounds for 7–13 months. Histological examination of the lungs revealed no signs of injury attributable to uranium exposure. In chronic-duration exposure tests, no histological damage attributable to uranium exposure to the lungs was observed. There was an absence of any other type of histological damage outside the kidneys (Cross et al. 1981a, 1981b; Stokinger et al. 1985). Dogs exposed to 15 mg/m3 of carnotite ore dust containing 0.6 mg U/m3 with a particle size activity median aerodynamic diameter (AMAD) of 1.5–2.1 μm for 1–4 years, 5 days a week, 4 hours a day, showed very slightly increased pulmonary resistance, which may not have been statistically significant. Histological findings included vesicular emphysema, which was present to a lesser degree in control animals. Fibrosis was not noted at this concentration (Cross et al. 1981a, 1982).
Exposure of 200 rats, 110 dogs, and 25 monkeys to 5 mg U/m3 as uranium dioxide dust for 1–5 years for
5.4 hours a day, 5 days a week did not result in histological damage in the lungs of the dogs or rats. Minimal patchy hyaline fibrosis was occasionally seen in the tracheobronchial lymph nodes of dogs and monkeys exposed for more than 3 years. No atypical epithelial changes were noted (Leach et al. 1970).
Because particles containing insoluble uranium compounds can reside in the lung for years, it is likely that radiotoxicity as well as chemical toxicity can result from inhalation exposure to highly enriched uranium compounds. Radiation effects on tissues from the alveolar regions of the lungs were examined in Albino HMT (Fischer 344) male rats exposed, nose-only, for 100 minutes to an aerosol of to 92.8% 235U-enriched uranium dioxide with a concentration of 2,273 nCi/m3 (84.1 kBq/m3) to 5,458 nCi/m3 (202 kBq/m3). Increases in the sizes and numbers of lung macrophages and type II3 cells, the numbers of
3Type I cells are alveolar lining cells that are involved with the transfer of substances from the alveolus through the wall to the blood. Type II cells are alveolar cells with two functions: oxidative enzymes for lung metabolism, and the production and secretion of the surfactant coating the alveolar surface.
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macrophages and type I cells, and a significant increase in the size of lysosomal granules within the macrophages were reported 8 days postexposure. At 7 days postexposure, 35 of the rats were further exposed to thermalized neutrons at a fluence of 1.0×1012 neutrons/cm2 over 2.5 minutes in order to study the combined effects of radiation and chemical toxicity. The radiation dose due to the neutrons and the fission fragments was about 600 rads, which is about 300 times greater than the radiation dose from the uranium dioxide alpha particles. No significant difference was found between the uranium dioxide-only group and those that were subsequently irradiated with neutrons, indicating that the extra radiation exposure caused no immediate pulmonary cellular reaction above that produced by uranium dioxide alone. This finding implies that the observed acute pulmonary effects were due to the metallotoxicity of the uranium dioxide rather than to the alpha radiation from the uranium (Morris et al. 1989). General damage to pulmonary structures, usually noncancerous alveolar epithelium damage of type II cells, can occur upon inhalation of insoluble reactive chemicals such as uranium salts and oxides. The main responses of epithelial cells to chronic injury are hyperplasia, hypertrophy, and transdifferentiation (metaplasia). These changes occur predominantly in proximal acinar regions where chronic injury often causes persistent lining of alveolar spaces by enlarged cuboidal cells that are derived from pre-existing type II cells, nonciliated epithelial cells from adjacent bronchioles, or a mixture of the two.
There is evidence that exposure to highly enriched uranium through inhaled or intratracheally instilled enriched uranium compounds adversely affect the epithelium of the lungs. Severe alveolar fibrosis or metaplasia was found in 72% of the sampled lung tissues from Fischer 344 rats exposed for 100 minutes to an aerosol of 92.8% enriched uranium dioxide at a radioactivity concentration of 5 μCi/m3 (137 kBq/m3) (-150 mg U/m3) to 10 μCi/m3 (270 kBq/m3) (-300 mg U/m3). Extensive lung disease of an unspecified nature was observed only in animals sacrificed at 720 days postexposure. The radioactivity concentration of the mixture was estimated as 1.91 kBq/g (51.6 nCi/mg), and the AMAD of the particles ranged from 2.7 to 3.2 μm (Morris et al. 1990).
In other animal studies, changes suggestive of damage from either radiation or diverse inorganic dust (fibrosis) were reported in lungs and tracheobronchial lymph nodes in Rhesus monkeys exposed by inhalation to 5.1 mg/m3 (as uranium dioxide) corresponding to a radioactivity concentration of 3.4 nCi/m3 (126 Bq/m3) for periods >3 years. Estimated cumulative alpha-radiation tissue doses were >500 rads (5 Gy) for the lungs and 7,000 rads (70 Gy) for the lymph nodes. Similarly exposed dogs also developed slight interstitial and vascular fibrosis of the lungs at lung alpha-radiation tissue doses of 760–1,280 rads (7.6–12.8 Gy) (Leach et al. 1970). The effect on the tracheobronchial lymph nodes in animals exposed for
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an additional 2 years ranged from involvement of a single node to complete destruction of all nodes, was dose-dependent, and showed a similarity to changes seen after inhalation exposure to plutonium as 238,239Pu dioxide (Leach et al. 1973). Renal damage was not observed in either dogs or monkeys, but fibrosis was found in monkey lung and both necrosis and fibrosis were found in dog and monkey lymph nodes. It was not clear whether the damage was chemically or radiologically induced, but the magnitude of the radiation doses and the presence of lung and lymph node damage in the absence of renal effects was suggestive to the authors of long-term radiation damage (Leach et al. 1970). However, such degenerative changes in the lungs have also been observed following prolonged exposure to diverse inorganic dust.
For more information about lung effects from plutonium and a review of the hazards associated with alpha-emitting radionuclide exposure, see the ATSDR Toxicological Profile for Plutonium (ATSDR 1990e) or Appendix D of this profile.
Cardiovascular Effects. No cardiovascular effects have been reported in humans after inhalation exposure to uranium. No effect on blood pressure or pulse rate was observed in a man accidentally exposed to powdered uranium tetrafluoride for 5 minutes (Zhao and Zhao 1990). Air concentration and mean particle size of the powder were not determined. Electrocardiograms and chest X-rays were normal shortly after the accident and over a 7.5-year follow-up period.
No cardiovascular effects were seen in rats exposed to 0.2 mg U/m3 (0.13 nCi U/m3) as uranium hexafluoride for 1 year (Stokinger et al. 1953) or in rats, mice, guinea pigs, and rabbits exposed to
4.8 mg U/m3 (3.2 nCi U/m3) triuranium octaoxide for 26 days (Dygert 1949c).
Gastrointestinal Effects. Inhalation exposure to uranium has generally not resulted in gastrointestinal effects in humans although transient effects occurred after one accidental exposure (Zhao and Zhao 1990). On the sixth day after a male worker at a uranium-enrichment plant was accidentally exposed for about 5 minutes in a closed room by inhalation to a high concentration of uranium tetrafluoride (natural uranium) powder, the patient reported nausea and loss of appetite. Air concentration and mean particle size of the powder were not determined. On post-accident day 8, the clinical findings were loss of appetite, abdominal pain, diarrhea, tenesmus, and pus and blood in the stool. On post-accident day 9, all parameters returned to normal. The study gave no indication of particle size for assessing deposition in the upper lung and no indication of whether fecal uranium analysis was undertaken to determine if the noted effects may have
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been mediated by the mucocilliary clearance of the uranium tetrafluoride from the lung and its subsequent swallowing to the gastrointestinal tract in accordance with the current ICRP lung model (ICRP 1994) or whether the signs were the result of another intestinal irritant. Gastrointestinal symptoms were not among the clinical signs reported for other workers accidentally exposed to uranium hexafluoride (Eisenbud and Quigley 1955; Moore and Kathren 1985; USNRC 1986).
Dogs, but not other species, appear susceptible to gastrointestinal effects after inhalation exposure to high concentrations of uranium compounds. Vomiting was observed during intermediate-duration exposure to
9.5 mg U/m3 uranyl nitrate (Roberts 1949), 18 mg U/m3 uranium tetrafluoride (Dygert 1949a), and to 9.2 mg U/m3 uranyl fluoride (Rothstein 1949a), It is possible that irritation of the gastrointestinal tract occurred either from clearance of uranium particles from the lungs or ingestion of uranium during these whole-body exposures. Histopathological examination of rat gastrointestinal tissues revealed no changes after 1-year exposures to 0.2 mg U/m3 uranium hexafluoride or uranium tetrachloride (Stockinger 1953).
Hematological Effects. Inhalation exposure to uranium compounds has generally had no effect, or only minor effects on hematological parameters in both humans and animals. In human studies, no hematological effects were found in a man accidentally exposed to powdered uranium tetrafluoride for 5 minutes (Zhao and Zhao 1990). Air concentration and mean particle size of the powder were not determined. Small but significant decreases in the hemoglobin concentration and the mean corpuscular hemoglobin concentration and significant increases in red blood cells counts and mean corpuscular volume were found in uranium miners who had worked for <5–20 years. All values measured were well within the normal range, such that values for individual miners could not be used as an estimate of exposure. No evidence of damage to red blood cell formation was found. The ambient concentration to which these workers had been exposed was not provided in the study (Vich and Kriklava 1970).
A study on the mortality among uranium mill workers found four deaths from lymphatic and hematopoietic tissue effects other than leukemia, while only one was statistically expected among these workers, who were occupationally exposed to uranium dust at airborne levels corresponding to a radioactivity concentration of 0.07 nCi/m3 (0.1 mg/m3). However, the authors of this study suggest that this excess may be due to irradiation of the lymph nodes by thorium-230 (230Th) (Archer et al. 1973b). No changes in hematological parameters were observed in humans occupationally exposed to uranium dust at a level of 1.7 nCi/m3 (63 Bq/m3 or 2.5 mg/m3) for 5 years (Eisenbud and Quigley 1955).
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Some intermediate-duration animal studies observed a range of hematological changes. Rats exposed to dusts of ammonium diuranate containing 6.8 mg U/m3 for 6 hours a day for 30 days showed a decrease of 1 million in red blood cell counts and a loss of 4 g of hemoglobin/100 mL of blood (Dygert 1949b). It was not stated whether exposure was for 30 consecutive days or on weekdays only. Rats exposed to airborne uranyl nitrate hexahydrate containing 9.5 mg U/m3 for 8 hours a day, 5 days a week for 30 exposure days showed decreased numbers of erythrocytes and hemoglobin (measured at 24 hours postexposure and weekly thereafter) (Roberts 1949). Increased percentages of lymphoid cells and myeloblasts in bone marrow were reported at termination in rats exposed to airborne uranium peroxide containing 15.4 mg U/m3 5 hours a day 5 days a week for 23 days (Dygert 1949d). A 4-week study in rats exposed to airborne uranium as uranium trioxide at a concentration corresponding to 16 mg U/m3 6 hours a day 6 days a week reported similar findings (significant increases in myeloblasts and lymphoid cells of bone marrow) (Rothstein 1949c). Rabbits and rats exposed to airborne uranium at a level corresponding to a uranium concentration of 0.13 mg/m3 as uranyl nitrate hexahydrate for 30 days exhibited altered blood function as indicated by decreased fibrinogen during the final week of exposure (Roberts 1949).
In contrast to the above findings, most other intermediate-duration animal inhalation studies with soluble and insoluble uranium compounds found no adverse effects on the blood. In intermediate-duration dosing studies lasting 23–40 days, inhalation exposure to various uranium compounds at the following concentrations produced no harmful effects on hematological parameters: 22 mg U/m3 as high-grade carnotite uranium ore to rats; 2.8 mg U/m3 as uranium dioxide or triuranium octaoxide to dogs; 22 mg U/m3 as uranium dioxide or triuranium octaoxide to rabbits; 11 mg U/m3 as uranium tetrachloride to rats; 2 mg U/m3 as uranium tetrachloride to rabbits; 1 mg U/m3 as uranium tetrachloride to dogs;
1 mg U/m3 as uranium hexafluoride to rabbits and dogs; 0.1 mg U/m3 as uranium hexafluoride to dogs;
2 mg U/m3 as triuranium octaoxide to mice; 14.5 mg U/m3 as uranium dioxide or triuranium octaoxide to rabbits; 14.5 mg U/m3 as triuranium octaoxide to guinea pigs and rabbits; 15.4 mg U/m3 as uranium peroxide to dogs, rabbits, and cats; or 4.8 mg/m3 as triuranium octaoxide to rats, mice, guinea pigs, and rabbits (Dygert 1949c, 1949d; Pozzani 1949; Rothermel 1949; Spiegl 1949).

In other intermediate-duration exposure studies, inhalation exposures to uranium dioxide dusts containing 1 mg U/m3 for 30 weeks and 2 mg U/m3 for 26 weeks in rabbits and guinea pigs, respectively (Stokinger et al. 1953), 19.4 mg U/m3 for 5 weeks in mice, and 9.2 mg U/m3 for 5 weeks in dogs and rats had no adverse effects on hematological parameters (Rothstein 1949b). Similarly, exposures to 9.2 mg U/m3 for
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5 weeks to rats and dogs (Rothstein 1949a); 16 mg U/m3 for 4 weeks to rats, rabbits, cats, and dogs (Rothstein 1949c); and 15 mg U/m3 as sodium diuranate to rats had no harmful effects on hematological parameters (Rothstein 1949d).
In chronic-duration exposures, dogs exposed to an airborne uranium concentration corresponding to a concentration of 0.2 mg U/m3 as uranium hexafluoride for 1 year exhibited a lengthening in blood clotting time with a decrease in blood fibrinogen levels (Stokinger et al. 1953). However, hamsters exposed to airborne carnotite uranium ore dust containing 0.7 mg U/m3 for 16–27 months exhibited no adverse hematological effects (Cross et al. 1981b). Similarly, no changes in hematological parameters were observed in rats, dogs, rabbits, and monkeys exposed to airborne uranium at concentrations ranging from 1 to 5.1 mg U/m3 for 1–5 years (Leach et al. 1970, 1973; Rothstein 1949b; Stokinger et al. 1953).
Musculoskeletal Effects. No studies were located regarding the chemical or radiation effects of uranium on the musculoskeletal system in humans or animals following inhalation exposure for any duration.
Hepatic Effects. No hepatic effects were found in a man accidentally exposed to powdered uranium tetrafluoride for 5 minutes (Zhao and Zhao 1990). Air concentration and mean particle size of the powder were not determined. Serum hepatic enzyme levels and liver function tests were within normal limits from the time of the incident through a 3-year follow-up period
Data from the available studies provide equivocal evidence that exposure of animals to uranium has effects on the liver, although the etiology for this effect is not clear. Urinary catalase, a measure of hepatic injury, was significantly increased in rabbits at an inhalation concentration of 0.13 mg U/m3 8 hours a day, 5 days a week for 30 exposure days (Roberts 1949). A slight decrease in hepatic lactate content was observed in rabbits following exposure to 15 mg U/m3 as sodium diuranate dust (Rothstein 1949d). Rabbits exposed to an inhalation concentration of 16 mg U/m3 as uranium trioxide dust for 4 weeks suffered moderate fatty livers in 63% of the animals that died (Rothstein 1949c). Focal necrosis of the liver was observed in rats exposed to an inhalation concentration of 0.4 mg U/m3 as uranium tetrafluoride for 30 days (Dygert 1949a). In other studies, no changes were found in the liver morphology, histology, or function in the following animals: rabbits exposed to 0.15 or 2 mg U/m3 as uranyl nitrate hexahydrate for 26 weeks; rats exposed to 14.5 mg U/m3 as triuranium octaoxide dust for
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26 days; rats exposed to 16 mg U/m3 as uranium trioxide for 4 weeks; mice and guinea pigs exposed to 3 mg U/m3 as high-grade uranium ore dust for 30 days; and rabbits exposed for 30 days to 22 mg U/m3 as high-grade uranium ore dust (contains uranium dioxide, triuranium octaoxide, and other potentially toxic contaminants) (Dygert 1949c; Pozzani 1949; Rothstein 1949c; Stokinger et al. 1953).
In chronic-duration exposure studies with animals, an unspecified strain of dogs exposed to ambient air concentrations of 0.05–0.2 mg U/m3 as uranium hexafluoride for 1 year exhibited increased and persistent bromosulfalein retention, indicative of impaired biliary function, at the 0.2 mg U/m3 concentration level (Stokinger et al. 1953).
Renal Effects. Uranium has been identified as a nephrotoxic metal, exerting its toxic effect by chemical action mostly in the proximal tubules in humans and animals. However, uranium is a less potent nephrotoxin than the classical nephrotoxic metals (cadmium, lead, mercury) (Goodman 1985). Many of the non-radioactive heavy metals such as lead, cadmium, arsenic, and mercury would produce very severe, perhaps fatal, injury at the levels of exposures reported for uranium in the literature (especially for miners and millers). The negative findings regarding renal injury among workers exposed to insoluble compounds are particularly significant in view of the high levels of exposure reported (Eisenbud and Quigley 1955). The United Nations Scientific Committee on the Effects of Atomic Radiation (UNSCEAR) has considered that limits for natural (and depleted) uranium in drinking water (the most important source of human exposure) should be based on the chemical toxicity rather than on a hypothetical radiological toxicity in skeletal tissues, which has not been observed in either people or animals (UNSCEAR 1993; Wrenn et al. 1985). However, it has been suggested that the renal damage from exposure to high-LET alpha-emitting heavy metals, such as uranium, may be the complementary effect of both the chemical toxicity and the radiotoxicity of these metals (Wrenn et al. 1987).
Several epidemiologic studies have found no increased mortality in uranium workers due to renal disease (Archer et al. 1973a, 1973b; Brown and Bloom 1987; Checkoway et al. 1988; Polednak and Frome 1981). Also, case studies showed that workers accidentally exposed to high levels of uranium did not suffer renal damage, even up to 38 years postexposure (Eisenbud and Quigley 1956; Kathren and Moore 1986), although the tests for renal damage used in these studies were not very sensitive. A recent comparison of kidney tissue obtained at autopsy from 7 uranium workers and 6 referents with no known exposure to uranium showed that the groups were indistinguishable by pathologists experienced in uranium-induced
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renal pathology (Russell et al. 1996). Three of 7 workers and 4 of 6 referents were categorized as abnormal. Uranium levels in the workers kidney tissue (estimated by alpha particle emission) ranged from 0.4 μg/kg to 249 μg/kg. One study on the kidney function of uranium mill workers chronically exposed to insoluble uranium (uranium dioxide) revealed renal tubular dysfunction as manifested by mild proteinuria, aminoaciduria, and a concentration-related clearance of β2-microglobulin relative to that of creatinine when compared to a referent group of cement workers. Air levels of uranium dioxide were not reported. The incidence and severity of these nephrotoxic signs correlated with the length of time that the uranium workers had spent in the area where insoluble uranium oxide yellowcake was dried and packaged (Saccomanno et al. 1982; Thun et al. 1985), which is typically the second dustiest area of the uranium mill following the ore crushing and grinding station. The data from this study are indicative of reduced reabsorption in the proximal renal tubules.
Delayed renal effects were observed after a male worker at a uranium enrichment plant was accidentally exposed to a high concentration of uranium tetrafluoride powder for about 5 minutes in a closed room. While renal parameters were normal during an initial 30-day observation period, the patient showed signs of nephrotoxicity beginning at post-accident day 68 as indicated by significantly elevated levels of urinary proteins, nonprotein nitrogen, amino acid nitrogen/creatinine, and decreased phenolsulfonpthalein excretion rate. These abnormalities persisted through day 1,065 but gradually returned to normal values (Zhao and Zhao 1990). The authors used uranium urinalysis data and a pharmacokinetic model (ICRP 1979) to estimate a kidney dose of 2.6 μg U/g kidney on post-accident day 1.
Renal effects were not observed in another accidental exposure (Fisher et al. 1990) in which 24 of 31 initially exposed workers were followed for 2 years. Estimated airborne concentrations were 20 mg uranium hexafluoride/m3 for a 1-minute exposure and 120 mg uranium hexafluoride/m3 for a 60-minute exposure (15.2 and 91 mg U/m3, respectively) (USNRC 1986). Initial intakes of workers involved in the accident were estimated from uranium excretion data and ranged from 470–24,000 μg uranium. Maximum uranium concentrations in the kidney were estimated by a kinetic model to be 0.048–2.5 μg U/g tissue (Fisher et al. 1991).
The pathogenesis of the kidney damage in animals indicates that regeneration of the tubular epithelium occurs upon discontinuation of exposure to uranium (Bentley et al. 1985; Dygert 1949b; Maynard and Hodge 1949; Pozzani 1949; Rothermel 1949; Rothstein 1949b, 1949c; Spiegl 1949; Stokinger et al. 1953).
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The magnitude of uranium intake that causes kidney damage depends on the type of uranium compound to which the animal has been exposed, appearing to depend on its solubility and oxidation state. For example, in dogs and monkeys, exposure to 5 mg U/m3 as uranium dioxide (insoluble) dust for up to 5 years produced no damage to the kidneys, even 6.5 years after the exposure ceased (Leach et al. 1970, 1973). Similarly, rats and guinea pigs were exposed to #10 mg U/m3 as uranium dioxide for 1 year without noticeable kidney pathology (Stokinger et al. 1953). Uranium dioxide is relatively insoluble in water and is retained in the lungs longer than the other more soluble uranium compounds (uranium tetrafluoride, uranyl fluoride, uranium tetrachloride, uranium peroxide, uranyl acetate, and uranyl nitrate hexahydrate), thereby causing higher toxicity to the lungs and lower toxicity to distal organs such as the kidney. In contrast, relatively soluble uranium compounds have been shown to cause renal tubular damage in dogs, guinea pigs, rabbits, and rats (Leach et al. 1984; Morrow et al. 1982; Roberts 1949; Stokinger et al. 1953). Apparently, the difference in effect is due to the extent of absorption of uranium deposited in the lungs and, thus, the fraction that eventually gets into the blood. Differences in species susceptibility have also been suggested to be an additional factor.
Renal effects can be produced in animals after acute-duration inhalation exposures to uranium. A 10-minute exposure to 637 mg U/m3 as uranium hexafluoride produced severe degeneration of the cortical tubules 5–8 days later in rats (Spiegl, 1949). These same effects were observed in dogs 1–3 days after a 1-hour exposure to 250 mg U/m3 as uranyl fluoride (Morrow et al. 1982). Proteinuria and glucosuria were also observed in rats after 2–10-minute exposures to uranium hexafluoride (Leach et al. 1984).
In intermediate-duration studies with guinea pigs, mice, rats, cats, rabbits, and dogs, inhalation exposures to a variety of uranium compounds were damaging to the kidneys. The effects were compound- and concentration-dependent and ranged from minimal microscopic lesions in tubular epithelium, increased urinary catalase, decreased diodrast (iodopyracet) clearance, and transient increased bromosulfalein retention (for low concentrations) to severe necrosis of the tubular epithelium (for high concentrations) in several species (Dygert 1949a, 1949b, 1949c; Pozzani 1949; Roberts 1949; Rothermel 1949; Rothstein 1949a, 1949c, 1949d; Spiegl 1949; Stokinger et al. 1953). In one of these intermediate-duration inhalation exposure studies, mice were exposed to uranium tetrachloride dust at ambient air concentrations of 0.1, 2.1, or 11 mg U/m3 for 3–7 hours a day 6 days a week for approximately 30 days. The exposure resulted in severe degeneration and necrosis of the renal-cortical tubular epithelium, and mortality, in the 11 mg U/m3 group by the third day. At the end of the study, moderate tubular
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degeneration was observed in the 2.1 mg U/m3 group and minimal degeneration in the 0.1 mg U/m3 group. (Rothermel 1949). In another intermediate-duration study, rats suffered renal injury (of inconsistent severity), which became apparent on or about the 7th day and pronounced by the 25th or 26th day, following inhalation exposure to uranyl nitrate hexahydrate at 0.13, 0.2, 0.9, 2.1, or 9.5 mg U/m3 daily for 8 hours per day, 5 days a week for 30 days. At 0.9 mg U/m3, the rats showed significant degenerative changes only in the renal tubules and no changes to the glomeruli. Rats exposed to
0.2 mg U/m3 exhibited only slight damage to the tubular epithelium of the kidneys. At 0.13 mg U/m3, slight renal tubular degeneration was observed in 1 of the 3 animals sacrificed after 28 days of exposure. Except for the group receiving no dietary supplement, no significant difference in blood CO2 values was seen at 14 days of exposure to uranium. Thirty days after the start of exposure, all groups exhibited increased blood nonprotein nitrogen (NPN) levels over 14 day values (maximum 111 mg/mL blood for the unsupplemented diet group). No clinical signs of toxicity were observed at any concentration level (Roberts 1949).
Dogs (of both sexes) exposed to 0.13 mg U/m3 as uranyl nitrate hexahydrate showed mild inner cortex changes after 10 days of exposure. The dogs were given full body exposures to aerosols with an AMAD assumed to be 1.5–2.1 μm; the average was 1.8 μm (Pozzani 1949). Severe nephritis masked any damage from uranium in one dog sacrificed after 10 days of exposure. The dogs showed a transient elevation in protein excretion between days 9 and 12 of exposure. Increased bromosulfalein retention was observed during the second and fourth weeks of exposure. No alterations to blood NPN or total blood CO2 were observed. Chloride clearance values, which were initially elevated and then became depressed in one dog, returned to normal 37 days after the beginning of exposure. Catalase and protein excretion increased significantly but returned to normal at the end of exposure. No significant changes in diodrast clearance, inulin clearance, and blood NPN levels were observed. Dogs exposed to 0.9 mg U/m3 exhibited mild inner cortex and medullary ray degeneration and necrosis with moderate epithelial regeneration. Two of the four showed a steady rise in NPN from the beginning of the experiment until they were sacrificed 12 days later, at which time NPN values were 252 and 356 mg%, respectively. Urinary protein in the dogs significantly increased between the 5th and 24th days. The dogs showed a decrease in inulin clearance during the third week of exposure, with a return toward normal values during the fifth week. There was decreased diodrast clearance throughout the observation period, indicating a severe derangement of the excretory capability for diodrast after 1 week (one dog showed a decrease of 69%). Diodrast clearances returned to normal by days 35–37. Two dogs showed a transient decrease in inulin clearance during the third week, lasting until the fifth week. All four dogs showed a drop in total
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blood CO2, attaining a minimum value between the first and seventh days. The minimum value was generally less than half that of controls, indicating severe acidosis. Glucose tolerance was significantly decreased. Large quantities of protein (400–800 mg%) and sugar were excreted. The greatest excretion occurred during the first 6 days of exposure and decreased thereafter. There was also a decrease in urinary creatinine excretion during the exposure. At the 2 mg U/m3 exposure level, the dogs did not show highly elevated NPN and blood urea nitrogen (BUN) values during exposure. There were no increases in blood NPN or BUN. All dogs exposed to 9.5 mg U/m3 had severe renal tubular damage. Four dogs showed renal injury followed by repair when they were sacrificed at the end of the exposure (Roberts 1949).
No treatment-related renal effects were seen in other studies when animals were exposed to uranium compounds by inhalation at concentrations as high as 10 mg U/m3 (as uranium dioxide) in guinea pigs for 28 weeks, 2 mg U/m3 (as uranyl nitrate hexahydrate) in guinea pigs for 26 weeks, and 16 mg U/m3 (as uranyl nitrate hexahydrate) in rats for 4 weeks (Rothstein 1949c; Stokinger et al. 1953).
The nephrotoxic effects of uranium in animals may also include damage to the glomerulus as evidenced by histopathological signs in the kidneys of rats and rabbits exposed to 15.4 mg U/m3 as uranium dioxide for 23 days (Dygert 1949d) and of dogs exposed to 15 mg U/m3 as uranyl fluoride for 5 weeks (Rothstein 1949d) and to 16 mg U/m3 as uranium trioxide for 4 weeks (Rothstein 1949c).
In chronic-duration inhalation studies with rats and dogs, uranium (as uranium tetrachloride, uranium tetrafluoride, uranyl nitrate hexahydrate, or uranium dioxide) exposures as low as 0.05 mg U/m3 and as high as 10 mg U/m3 for 1–5 years were damaging to the kidneys. Nephrotoxic effects found in these animals ranged from minimal microscopic lesions in tubular epithelium (for low concentrations) to acute tubular necrosis (for high concentrations) (Leach et al. 1970; Stokinger et al. 1953). In one of these chronic-duration studies, dogs were exposed to ambient air concentrations of 0.05 or 0.2 mg U/m3 as uranium hexafluoride for 1 year for a total of 1,680 exposure hours. The UF6 was rapidly hydrolyzed to HF gas and UO2F2 fumes, whose AMAD was 0.1 μm. After 10 days in the study, there was evidence of mild tubular injury, which was characterized by desquamation of the epithelium and active regeneration in the proximal convoluted tubule in the inner cortex of the kidneys in 86% of animals exposed to
0.2 mg U/m3. From the 16th week to the end of the study, regeneration of the tubular epithelium was almost complete, with a few flattened atrophic tubules in the inner zone of the cortex. These mild nephrotoxic effects were also observed in 12% of the 0.05 mg U/m3 exposed animals. Blood non-protein
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nitrogen (NPN) levels were normal (elevated blood NPN levels indicate a decrease in renal filtration capacity, similarly to elevated blood urea nitrogen (BUN)). Observed changes in urinary protein were inconsistent and insignificant (Stokinger et al. 1953).
In another study, dogs of both sexes (9–12 M, 9–13 F) were exposed to concentrations of 0.04, 0.15, 0.25, or 2 mg U/m3 as uranyl nitrate for 6 hours a day, 5.5 days a week for 1 year. The AMAD of the aerosols was given as 2–5 μm. At the termination of the study, histological and biochemical examinations revealed minimal microscopic lesions in the renal tubules and transient increases in blood NPN in the
0.25 mg U/m3 concentration-level dogs. Transient increases in blood NPN were also observed at higher concentration levels. There were transient decreases in plasma CO2, although liver function was normal. No significant weight loss was observed in the dogs (Stokinger et al. 1953).
No treatment-related renal effects were seen when Rhesus monkeys and dogs were exposed to uranium dioxide by inhalation at airborne concentrations as high as 5.1 mg U/m3 for 1–5 years (Leach et al. 1973). Blood NPN levels were consistently elevated in Rhesus monkeys although no renal histopathology was evident (Leach et al. 1973).
Endocrine Effects. A single study was found that reported on possible effects of uranium on the endocrine system. In this study, no histopathology was seen in the endocrine organs (adrenal, pancreas) in rats given 0.2 mg U/m3 as uranium tetrachloride for 1 year (Stokinger et al. 1953).
Dermal Effects. No dermal effects were found in a man accidentally exposed to powdered uranium tetrafluoride for 5 minutes (Zhao and Zhao 1990). Histopathologic examination of the skin was normal in rats exposed to 0.2 mg U/m3 as uranium tetrachloride for 1 year (Stokinger et al. 1953).
Ocular Effects. Chemical burns to the eyes were reported in humans after accidental exposure to uranium hexafluoride (Kathren and Moore 1986). Conjunctivitis and eye irritation have also been reported in animals after exposure to uranium hexafluoride (Spiegl 1949) and to uranium tetrachloride (Dygert 1949a). Ocular effects were due to direct contact of the eye with vapor or aerosols.
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Body Weight Effects. In general, inhalation of insoluble uranium compounds did not significantly affect body weight in animals. Decreased body weight was observed with the more water-soluble compounds. A 30% decrease in body weight was reported for rabbits exposed to 11 mg U/m3 as uranium tetrachloride dust for 35–40 days. Mice and guinea pigs experienced unspecified weight loss and 13% weight loss, respectively, following exposure to 13 mg U/m3 as uranium hexafluoride for 30 days. Rabbits suffered 12% weight loss following exposure to 0.2 mg U/m3 as airborne uranium hexafluoride for 30 days (Spiegl 1949). Mild to severe weight loss was observed in several species during exposure to uranyl nitrate hexahydrate (Roberts 1949). Rabbits lost 22% of their body weight during a 30 day exposure to 0.9 mg U/m3, dogs and cats lost approximately 25% of their body weight during a similar exposure to 9.5 mg U/m3. Similar effects were observed with uranium tetrafluoride (Dygert 1949a). Rabbits, rat, cats, and dogs all experienced a greater than 20% weight loss during 30 days exposure to 18 mg U/m3.
Several intermediate-duration animal inhalation studies with soluble and insoluble uranium compounds found no significant adverse effects on body weight. In short-term intermediate-duration dosing studies lasting from 23 to 40 days, exposure to concentrations at the following levels were without significant effects on body weight: 22 mg U/m3 as high-grade or carnotite uranium ore to rats, 2.9 mg U/m3 as uranium dioxide or triuranium octaoxide to dogs, 22 mg U/m3 as uranium dioxide or triuranium octaoxide to rabbits, 11 mg U/m3 as uranium tetrachloride to rats, 2.1 mg U/m3 as uranium tetrachloride to rabbits,
1.1 mg U/m3 as uranium tetrachloride to dogs, 13 mg U/m3 as uranium hexafluoride to rabbits and dogs,
0.2 mg U/m3 as uranium hexafluoride to dogs and guinea pigs, 14.5 mg U/m3 as triuranium octaoxide to mice, and 4.8 mg U/m3 as triuranium octaoxide to guinea pigs and rabbits (Dygert 1949c; Spiegl 1949); 15 mg U/m3 as uranium peroxide to cats and rabbits (Dygert 1949d); 15 mg U/m3 as carnotite ore (mostly uranium dioxide, triuranium octaoxide) to dogs or 22 mg U/m3 as carnotite ore to rabbits for 30 days (Pozzani 1949); and 1 mg U/m3 for 30 weeks to rabbits or 2 mg U/m3 for 26 weeks to rabbits and guinea pigs (Stokinger et al. 1953). Exposures of rats to 13 mg U/m3 or of rabbits to 0.1 mg U/m3 as uranium hexafluoride for 30 days also were without harmful effects (Spiegl 1949).
No effects on body weight were observed after several intermediate-duration dosing studies that lasted 4–5 weeks. These studies researched exposures by the inhalation route as follows: 16 mg U/m3 as uranium trioxide to rats, rabbits, dogs, and cats; 19 mg U/m3 as uranium dioxide to mice; 16 mg U/m3 as uranium dioxide to guinea pigs; 9.2 mg U/m3 as uranyl fluoride to dogs and rabbits; 2.2 mg U/m3 as uranyl fluoride to rats; 9.2 mg U/m3 as uranium dioxide to dogs; 19.2 mg U/m3 as uranium dioxide to rabbits; 15 mg U/m3
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as sodium diuranate to rats and dogs; and 12 mg U/m3 as ammonium diuranate to rats for 30 days (8 hours a day, 5 days a week for 6 weeks) (Rothstein 1949a, 1949b, 1949c, 1949d; Stokinger et al. 1953). Hamsters exposed to 0.8 mg U/m3 as carnotite uranium ore by inhalation for 16–27 months also exhibited no adverse body weight effects (Cross et al. 1981b). Similarly, no changes in body weight were observed in rats, dogs, rabbits, and monkeys exposed to airborne uranium dioxide at 0.1–5 mg U/m3 for 1–5 years (Leach et al. 1970, 1973; Stokinger et al. 1953).
In chronic-duration studies, exposure to inhalation concentrations of 3 mg U/m3 as uranium dioxide to monkeys for 5 years produced no significant body weight changes (Leach et al. 1970).
Other Systemic Effects. Several general effects have been attributed to uranium inhalation exposure. In animal studies, dogs exposed to 13 mg U/m3 as uranium hexafluoride for 30 days exhibited decreased water intake (Spiegl 1949). Reduced food intake was also observed in a 4-week study of rats and mice exposed to 16 mg U/m3 as uranium trioxide (Rothstein 1949c) and in a 5-week study of rats and mice exposed to 15 mg U/m3 as sodium diuranate for 6 hours per day, 51/2 days per week (Rothstein 1949d).
2.2.1.3 Immunological and Lymphoreticular Effects
Although no studies were located that specifically tested immunological effects in humans following inhalation exposure to uranium, all epidemiologic studies of workers in uranium mines and fuel fabrication plants showed no increased incidence of death due to diseases of the immune system (Brown and Bloom 1987; Checkoway et al. 1988; Keane and Polednak 1983; Polednak and Frome 1981).
Human studies that assessed damage to cellular immune components following inhalation exposure to uranium found no clear evidence of an immunotoxic potential for uranium. No association was found between the uranium exposure and the development of abnormal leukocytes in workers employed for 12–18 years at a nuclear fuels production facility (Cragle et al. 1988). Increases in the number of fatal malignant disease of the lymphatic and hematopoietic tissue reported among uranium mill workers may have been caused by other carcinogens in the work environment such as 230Th. The authors of this report estimated that the workers were exposed to 8–5,100 mg/m3 (median 110 mg/m3) uranium mill dust, which contains 230Th as a natural component (Archer et al. 1973b).
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In animal studies, rats exposed to dusts of ammonium diuranate containing 6.8 mg U/m3 for 6 hours a day, 5 days per week for 30 days developed a rise in neutrophils, a decrease in lymphocytes, a moderate fall in the white blood cell count, and a rise in the number of eosinophils (Dygert 1949b). Rats exposed to airborne uranyl nitrate hexahydrate containing 9.5 mg U/m3 8 hours a day, 5 days a week for 30 exposure days showed an initial increase and a subsequent decrease in the absolute number of lymphocytes and neutrophils (Roberts 1949). Focal necrosis of the spleen and edematous cecal lymph nodes were observed in some rats exposed for 30 days for 6 hours a day to 0.4 and 4 mg U/m3 uranium tertrafluoride (Dygert 1949a). However, these effects were not observed at 18 mg U/m3, so the significance of this finding is unclear.
No histopathological changes or accumulation of uranium were evident in the spleens of 110 dogs and 25 monkeys exposed to uranium dioxide dusts (5 mg U/m3) for 6 hours a day, 5 days a week, 1–5 years and then monitored for up to 6.5 more years. Similar results were seen for rats similarly exposed for 1 year (Leach et al. 1970, 1973). Rats, rabbits, guinea pigs, and dogs exposed to dusts of various uranium compounds for 7–12 months showed no significant histological changes in the lymph nodes and
marrow (Stokinger et al. 1953).
There is some evidence from animal studies that exposure to $90% enriched uranium may affect the immune system. Increased macrophage activity, associated with localized alpha tracks in all 5 lobes of the lungs, was seen in Fischer 344 rats exposed to 6,825.5 nCi/m3 (252 kBq/m3) through inhalation exposure to enriched uranium dioxide for 100 minutes. The increased activity was evident from days 1–7, 180, 360, 540, and 720 with increases in percent activity of 0.44, 2.15, 19.70, 6.54, and 37.84, respectively. The number and size of macrophage clusters in the lung increased with time postexposure. The radioactive material concentration of the mixture was estimated as 1.91 kBq/mg (51.6 nCi/mg) (Morris et al. 1992). The degree of enrichment was calculated based on this specific activity.
Albino HMT (Fischer 344) male rats were exposed to 92.8% enriched uranium dioxide with a concentration ranging from 2,274.2 nCi/m3 (84.1 kBq/m3) to 5,458 nCi/m3 (202 kBq/m3). Increases in the sizes and numbers of lung macrophages, with a significant increase in the size of lysosomal granules within the macrophages, were reported 8 days postexposure (Morris et al. 1989).
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Dogs exposed to airborne uranium dioxide concentrations of 5.1 mg/m3 for 1–5 years showed lymph node fibrosis in the lungs. Rhesus monkeys similarly exposed for 5 years showed fibrotic changes in the tracheobronchial lymph nodes. The investigators of these studies concluded that although these effects could not be extrapolated to humans because of the absence of squamous cell carcinomas in the lungs, the changes were suggestive of radiation injury (Leach et al. 1973). However, the morphological changes observed in these studies were similar to observations in humans and animals as a result of exposure to diverse inorganic dust (Dockery et al. 1993).
The highest NOAEL values and all reliable LOAEL values in each species and duration category for immunological effects from chemical exposures by the inhalation route to uranium are presented in Table 2-1 and plotted in Figure 2-1.
2.2.1.4 Neurological Effects
Uranium has not been shown to cause damage to the nervous system of humans by metallotoxic or radiotoxic action following inhalation exposures for any duration. Although no studies were located that specifically tested neurological effects in animals following inhalation exposure to uranium, none of the available studies reported any neurological deficits, such as narcosis, ataxia, or cholinergic signs. Clinical signs in humans following acute exposure to enriched uranium included dizziness and anorexia in one man 6 days after being exposed for 5 minutes to uranium tetrafluoride by inhalation (Zhao and Zhao 1990), but did not include neurological effects in others similarly exposed to uranium hexafluoride (Kathren and Moore 1986; USNRC 1986). Some of the victims were evaluated for as long as 38 years after exposure (Kathren and Moore 1986). In longer-term exposures, epidemiologic studies found no increase in deaths from brain tumors or other neurological diseases that could be attributed to uranium in workers at uranium-processing plants (Brown and Bloom 1987; Carpenter et al. 1988; Cragle et al. 1988; Polednak and Frome 1981; Reyes et al. 1984). The autopsy reports also did not reveal any other structural pathology of the central nervous system. In a retrospective study, more deaths than expected were found from central and peripheral nervous system diseases (SMR=2.98) in employees in a nuclear fuels fabrication plant. However, the employees were also concurrently exposed to other radiological and chemical agents. The investigators of this study concluded that there was no etiology associated with uranium for the central nervous system and peripheral nervous system diseases (Hadjimichael et al. 1983).
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In intermediate-duration animal studies, neurological signs were observed in dogs and cats following inhalation exposure to uranium. On the 13th day of a 30-day study, dogs exposed to 0.5, 3, 4, or 18 mg U/m3 as uranium hexafluoride gas by inhalation exhibited muscular weakness followed by instability of gait indicative of neurological dysfunction at the highest concentration tested (Dygert 1949a). Anorexia observed in another 8 hours a day, 5 days a week, 30-day study with dogs exposed to an inhalation concentration of 9.5 mg U/m3 as uranyl nitrate hexahydrate may also have had its origin in neurological dysfunction (Roberts 1949). Similarly, cats exposed to an inhalation concentration of 18 mg U/m3 as uranium tetrafluoride exhibited unsteady gait on the 7th day in a 30-day study (Dygert 1949a). In 5 week studies (8 hours a day, 5 days a week), dogs and cats exposed to 0.15, 2.2, or
9.2 mg U/m3 as uranyl fluoride suffered anorexia, severe muscle weakness, and lassitude at the highest concentration tested (Rothstein 1949a). These studies did not assess the potential implications of hydrofluoric acid and fluoride ion exposure.
The highest NOAEL values and all reliable LOAEL values in each species and duration category for neurological effects by the inhalation route to uranium are presented in Table 2-1 and plotted in Figure 2-1.
2.2.1.5 Reproductive Effects
It is unlikely that inhalation of uranium produces a significant effect on reproductive health. Studies of one human population group (miners) were located which identified a reproductive effect associated with the inhalation exposure of mine air, but the association with uranium compounds was unclear, and the other miner studies observed no reproductive effects. Also, no adverse animal studies were found.
Three studies of one mining population were located that equivocally associated reproductive effects in humans following inhalation exposure to uranium. The studies reported that male uranium miners were found to have more first-born female children than expected, suggesting that uranium’s alpha radiation damaged the y-chromosomes of the miners (Muller et al. 1967; Waxweiler et al. 1981b; Wiese 1981). In addition, it is not certain if the effect described is from exposure to uranium because the workers were also exposed to 222Rn, chlorine, hydrofluoric acid, lead sulfate, nickel, nitric acid and nitrogen oxides, silicon dioxide, and sulfuric acid (Dupree et al. 1987).
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No animal studies were located that described reproductive effects following inhalation exposure to uranium for any duration of exposure.
2.2.1.6 Developmental Effects
No studies were located which reported effects of uranium on development in humans or animals following inhalation exposures for any duration. The Department of Defense has preliminarily evaluated developmental effects among service members who were actually or potentially exposed to depleted uranium.
2.2.1.7 Genotoxic Effects
No information was located regarding the toxicity of uranium to genetic material in humans or animals following inhalation exposures for any duration.
In human studies, chromosome aberrations have been found in cultured lymphocytes of uranium miners. Miners who had more atypical bronchial cell cytology had more chromosomal aberrations, and some of the aberrations increased with increasing exposure to radon and its decay products. The investigators of the study concluded that this is probably a valid health risk indicator for miner groups, but that it has only limited applicability to individual miners (Brandom et al. 1978). In a similar study with uranium miners in Czechoslovakia, no increased incidence of aberrant DNA or chromosomes attributable to exposure to uranium was found. An increased occurrence of molds (genus Aspergillus and Penicillium) that produce mycotoxins was observed, suggesting that the inhaled dust was contaminated with these genotoxic microorganisms (Sram et al. 1993). A cytogenic study of men occupationally exposed to uranium found higher levels of chromosome aberrations in the miners than in controls. The investigators of this study concluded that this increase may be attributable to smoking (Martin et al. 1991). In addition, because the miners were also concurrently exposed to chlorine, hydrofluoric acid, lead sulfate, nickel, nitric acid and nitrogen oxides, silicon dioxide, diesel smoke, and sulfuric acid in addition to 222Rn, it is unlikely that the effects described in these studies were related in any way to exposure to uranium (Dupree et al. 1987). Other genotoxicity studies are discussed in Section 2.5.
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2.2.1.8 Cancer
The National Toxicology Program (NTP) has not evaluated uranium compounds in rodent cancer bioassays by any route for the potential to induce cancer in humans. However, because uranium emits predominantly high-LET alpha particles, current theories on gene mutation and apoptotic mechanisms of cancer promotion by high-LET alpha radiation suggest a concern for carcinogenesis from uranium’s radioactivity (BEIR 1980, 1988, 1990; Otake and Schull 1984; Sanders 1986; UNSCEAR 1982, 1986, 1988) (see Appendix D for a review of the hazards associated with radionuclide exposure).
Although several studies of uranium miners found increased deaths from lung cancer, it is difficult to attribute these cancers to uranium exposure because the miners were also concurrently exposed to known cancer-inducing agents (principally tobacco smoke, radon and its decay products, silica and other dusts, and diesel engine exhaust fumes) and the studies attributed the cancers to exposure to these toxicants and not to uranium exposure (Archer et al. 1973a; Auerbach et al. 1978; Band et al. 1980; Gottlieb and Husen 1982; Kusiak et al. 1993; Lundin et al. 1969; Saccomanno et al. 1971, 1976, 1986; Samet et al. 1984; Whittemore and McMillan 1983). Short-lived radon daughters alone, to which these miners were concurrently exposed, have been shown to increase the risk of developing lung cancer (Saccomanno et al. 1986). In addition, smoking appeared to increase the risk of developing lung cancer from exposure to radon daughters (Band et al. 1980). The available case-control or clinical studies of uranium-processing nuclear plant workers also generally report equivocal findings of cancer induction without establishing any uranium causality (Cookfair et al. 1983; Polednak and Frome 1981).
A review of the morphology of the tumor types induced in the lungs of rats and humans by radiation identified bronchoalveolar adenoma and bronchoalveolar carcinoma, papillary adenocarcinoma, squamous cell carcinoma, adenosquamous carcinoma, and hemangiosarcoma. All the tumor types originated from the alveolar parenchyma region of the lungs. Of these tumor types, squamous cell carcinomas are most often associated with radiation exposure. In irradiated rats, the squamous cell carcinomas are less well differentiated and decidedly more locally invasive. Although cystic squamous tumors do occur after irradiation, the wall of these tumors is less differentiated. The pathogenesis of the radiation-induced squamous tumors appeared to be different from that of chemically induced tumors. The one common feature of the two tumor types may be chronic injury to alveolar type II cells (Hahn 1989).
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Squamous cell metaplasia was the predominant aberrant cell type found in many of these cases. Squamous cell metaplasia is found in the young and old and does not always represent a benign-tomalignant process. More frequently, a nonspecific bronchial epithelial reaction develops, and this reaction is readily reversible with the disappearance of the toxic, infectious, or inflammatory factors that caused it. Although, squamous cell metaplasia may develop into neoplasia, patients with neoplasia also shed a variety of metaplastic squamous cells. In a study of 120 uranium miners who died from primary cancer of the lungs, squamous cell metaplasia progressed over time and developed into neoplasia of the lungs in 15–20 years (Saccomanno et al. 1976, 1982). However, a study that reviewed efforts to test uranium miners concluded that radon-progeny exposure may not cause any cell type of lung tumor other than the so-called small-cell (oat cell) carcinoma. The incidence of oat cell cancer of the lungs has decreased over the last 20 years and currently accounts for slightly more than 22% of developing neoplasia in uranium miners (Saccomanno et al. 1982).
An excess of lung cancers has been found in underground uranium miners from the Grants, New Mexico, area. Of 3,055 miners who worked for at least one year prior to 1971, a total of 58 died of lung cancer by the middle of 1985. Of the 43 cancers which had been examined histopathologically, 27 (63%) were small-cell, 14 (33%) were epidermoid, 1 (2%) was adenocarcinoma, and 1 (2%) was large-cell. These mortality data could not be related to the total radon exposure; radon exposure data for the individual miners was complete since 1967, but only mine-average concentrations had been determined for the period prior to that time (Samet et al. 1986) The radon concentration in mine air is measured in working levels (WL), where 1 WL = 100 pCi/L Rn in equilibrium with its daughters, and the total exposure to radon is measured in WLM s where 1 WLM = 170 WL, or the equivalent of breathing air at a concentration of 1 WL for a period of 170 hours (the typical miner work month). A total of 8,487 miners employed between 1948 and 1980 at the Beaver Edge uranium mine in Saskatchewan, Canada, exhibited significant increase in lung cancer deaths when compared to Canadian male mortality rates (65 in exposed populations as opposed to 34.2 expected [p<0.05]). A higher incidence of lung cancer was found in workers exposed to more radon than 5 WLM (46 observed as opposed to 15.8 expected) than those exposed to 0–4 WLM (19 observed as opposed to 18.7 expected). A significant relationship was found between radon exposure and increase of lung cancer (3.3% per WLM and 20.8% per WLM/106 person-years). The age at first exposure also had a significant effect on risk; those first exposed before the age of 30 were at lower risk than those first exposed at or after 30 years of age. The authors suggested that exposure to radon daughters was the major factor, and it may be a contributory factor to lung cancer in
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nonsmokers in the general population (Howe et al. 1986). The frequency of squamous cell cancer increased, relative to other types, with increased levels and durations of smoking, but relative frequency was not affected by radiation exposure. The relative frequencies of small-cell cancer and adenocarcinoma from radiation exposure were less affected by smoking (Archer et al. 1973a; Land et al. 1993; Saccomanno et al. 1988).
Histological examination of lungs from seven underground male uranium miners (ages 52–73) who had cancer and who also had been routinely exposed to radon daughters and other potential carcinogens in the mine environment showed elevated concentrations of 238U and 234U. Four of the seven lungs had squamous cell carcinoma, one had a carcinoma in the left upper lobe, one had carcinoma of the ascending colon, and one had carcinoma in situ in the lung. The average radiation dose from uranium was approximately 2 mrad/year (2×10-5 Gy/year) compared with more than the 360 mrad (3.6×10-3 Gy) dose to the typical U.S. resident from all sources of radiation. Five of the seven miners smoked at least half a pack of cigarettes per day (Wrenn et al. 1983).
A study of miners in northern Ontario with previous inhalation exposure to uranium dust at levels of 0–181 mg U/m3 (0–121 nCi/m3 [0–4,487 Bq/m3]) and a diagnosis of lung cancer found a linear relationship between uranium dose and incidence of lung cancer, but no relationship to uranium exposure was suggested. The latency period was shorter for those employed for a short period of time. Oat cell, anaplastic, small-cell tumors were found more often than squamous, large-cell, poorly differentiated tumors in workers exposed for a short time (Chovil and Chir 1981; Sanders 1986). The frequency of squamous cell cancer in U.S. uranium miners increased, relative to other types, with increased levels and durations of smoking; but relative frequency was not affected by radiation (presumed to be mostly from radon daughters) exposure, which indicated a more likely smoking etiology. However, the relative frequencies of small-cell cancer and adenocarcinoma were less affected by smoking history than by increasing radiation dose. The miners in one of these studies were exposed to a cumulative radiation dose from radon daughters of 40–9,700 WLM (Archer et al. 1973a; Land et al. 1993; Saccomanno et al. 1971, 1982, 1988). A reanalysis study in which sputum samples from 98,181 uranium miners employed on the Colorado Plateau between 1960 and 1980 were collected and used in a cytological analysis for the early detection of cancer development found a significant relationship between exposure to radon decay products and positive cytological diagnosis. No evidence was found linking lung cancer with exposure to uranium. No synergism was seen between age, smoking, and mining exposure, although an additive
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effect was seen. No increase in lung cancer was found in men exposed to radon for 1 nCi (37 Bq) (Reyes et al. 1984).
The available human studies that investigated the association between the development of bone sarcomas and exposure to uranium failed to produce evidence for the development of bone sarcomas or bone cancers of any type (Archer et al. 1973a; Chovil and Chir 1981; Cookfair et al. 1983; Cragle et al. 1988; Gottlieb and Husen 1982; Grace et al. 1980; Kusiak et al. 1993; Land et al. 1993; Polednak et al. 1982; Reyes et al. 1984; Saccomanno et al. 1971, 1976, 1988; Samet et al. 1986; Wrenn and Singh 1983).
Development of lymphatic malignancies (other than leukemia) has also been associated with inhalation exposure to materials associated with uranium. In a study of 2,002 uranium millers, 6 deaths from lymphatic malignancies occurred when 2.6 were expected. The latency period was 20 years (Waxweiler et al. 1983). Another study of uranium mill workers found a slight increase in deaths from tumors of the lymphatic and hematopoietic tissue (Archer et al. 1973b). The authors suggested that this finding might not be due to uranium itself, but rather due to irradiation of the lymph nodes by 230Th, a decay product of 234U and a member of the 238U decay chain.
In intermediate-duration animal studies, golden Syrian hamsters exposed to carnotite uranium ore dust (AMAD=1.5–2.1 μm) at a concentration of 19 mg U/m3 by inhalation for 16 months failed to shown signs of cancer development upon examination of selected tissues including lungs, trachea, liver, kidneys, spleen, heart, and any abnormal tissue. As compared to unexposed controls, the hamsters had significantly more necrotic liver foci and inflammatory lung responses (Cross et al. 1981b).
In the same study, the results of exposure of golden Syrian hamsters for 16–27 months to concentrations of radon progeny, uranium ore dust (0.5 nCi/m3 [18.5 Bq/m3]), or a combination of uranium and radon progeny provided evidence that, while prolonged exposure to uranium dust causes inflammation and
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proliferative pulmonary changes, inhalation of radon progeny produced bronchiolar epithelial hyperplasia and changes in the alveolar epithelium in hamsters. The authors also concluded that exposure to radon progeny and development of squamous metaplasia and carcinoma were related. The animals had cumulative radon progeny exposures of more than 8,000 WLM. Pulmonary neoplasms occurred in the three radon-progeny-exposed hamsters and in one hamster exposed to a combination of uranium, radon, and radon progeny. Both the hamsters exposed to radon progeny and those exposed to a combination of uranium and radon progeny had a significantly greater incidence of adenomatous proliferative changes in the alveolar epithelium. Uranium ore-exposed hamsters had significantly more necrotic liver foci and inflammatory lung responses than animals from other exposure groups. Specifically, one pheochromocytoma (zero in controls), one melanoma (zero in controls), one hemangioendothelioma (one in controls), two reticulum cell sarcomas (three in controls), and one adrenal cell carcinoma (zero in controls) were seen in animals exposed to uranium dust alone. Two osteosarcomas (zero in controls) were reported in animals exposed to the mixture of uranium ore dust and radon progeny. Four reticulum cell sarcomas (three in controls) and one adrenal cell sarcoma (zero in controls) were also seen in these animals. In animals exposed to radon progeny alone, one undifferentiated sarcoma (zero in controls), three reticulum cell sarcomas (three in controls), and one myelogenous leukemia (one in controls) were observed (Cross et al. 1981b).
In chronic animal studies, analysis of Beagle dogs exposed to 3.4 nCi/m3 (126 Bq/m3 or 5 mg U/m3) uranium dioxide found frank pulmonary neoplasms and atypical epithelial proliferation in 30–46% of the animals. The lung dose was estimated as 600–700 rads (6–7 Gy). Spontaneous tumors in dogs were infrequent, and the incidence found in this study was 50–100 times higher than the expected rate of spontaneous tumors. The authors of the study recommended against the extrapolation of these findings to humans because these glandular neoplasms do not occur frequently in humans (Leach et al. 1973).
Cancer effect levels (CELs) for chemical and radiation inhalation exposure to uranium are shown in Tables 2-1 and 2-2 and plotted in Figures 2-1 and 2-2.
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2.2.2 Oral Exposure
The oral toxicity of uranium compounds has been evaluated in several animal species. The maximal dosage just failing to be lethal for rats in a 30-day feeding test was about 0.5% uranium compound in the diet for the 3 soluble compounds (uranyl nitrate hexahydrate, uranyl tetrafluoride, and uranium tetrachloride) and 20% uranium compound for the 3 insoluble uranium compounds (uranium dioxide, uranium trioxide, and triuranium octaoxide) tested. Some of these studies sweetened the feed to make it edible. No amount of insoluble uranium compounds acceptable to the rat was lethal. Dietary levels of 1–4% soluble uranium compound produced 50% mortality in 30 days. The marked difference in the toxicity of soluble and insoluble uranium compounds is attributable to the ease of absorption and, thus, the dose that reaches the target organs. In general, the water-soluble compounds are more toxic by the oral route because of the greater ease of absorption in the gastrointestinal tract (Domingo et al. 1987, 1989a, 1989b; Goel et al. 1980; Maynard and Hodge 1949; Paternain et al. 1989). In a summary of the oral toxicity in both rats and dogs, several uranium compounds were ordered by relative toxicity as follows: very toxic compounds included uranium tetrachloride, uranium peroxide, and uranyl fluoride; toxic compounds included uranium nitrate hexahydrate, uranyl acetate, ammonium diuranate, sodium diuranate, uranium trioxide, and high-grade uranium ore (carnotite); practically nontoxic compounds were uranium tetrafluoride, triuranium octaoxide, and uranium dioxide (Maynard and Hodge 1949).
2.2.2.1 Death
There are no reports of human deaths from oral exposure to uranium compounds. However, data from animal studies demonstrate that soluble uranium compounds, at very high intake levels, can be lethal to animals through the oral route for all durations of exposure. Uranium compounds at these concentrations are not palatable to animals and require sweetening.
Oral LD50 (lethal dose, 50% mortality rate) values of 114 and 136 mg U/kg have been estimated for male Sprague-Dawley rats and male Swiss-Webster mice, respectively, following single gavage administrations of uranyl acetate dihydrate (Domingo et al. 1987). Mortality occurred in pregnant Swiss mice exposed to 0.028, 0.28, 2.8, 28 mg U/kg/day uranium as uranyl acetate dihydrate by gavage in water from day 13 of gestation through postnatal day 21. Two dams in the 2.8 and three in the 28 mg U/kg/day groups died before delivery (Domingo et al. 1989b). Deaths were also reported in mice during the first 10
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days of feeding studies with uranyl nitrate (8 of 25 at 925 mg U/kg/day) and with uranyl fluoride (2 of 25 at 452 mg/kg/day) (Tannenbaum and Silverstone 1951)
In 30-day oral studies, oral LD50 values for both sexes of rats of an unspecified strain given uranyl fluoride or uranyl nitrate hexahydrate have been estimated as 540 and 1,579 mg U/kg/day, respectively. Oral LD50 values were 658 and 1,096 mg U/kg/day as uranium tetrachloride for male and female rats, respectively, in a similar 30-day study (Maynard and Hodge 1949). Another 30-day study, in which male and female rats of an unspecified strain were exposed to oral uranium peroxide doses, oral LD50 values were estimated as 827 and 1,103 mg U/kg/day, respectively (Maynard and Hodge 1949). In other intermediate-duration feeding studies with rats, 16% mortality was reported in the animals following dietary administration of 664 mg U/kg/day for 30 days. Most of the animals died from complications of chemically induced kidney damage (Maynard et al. 1953).
Two-year feeding studies with uranyl fluoride, uranyl nitrate hexahydrate, uranium tetrafluoride, and uranium dioxide showed that chronic intake of large amounts of uranium can lead to a decrease in lifespan. The largest daily intake that did not affect longevity in the rat was 81 mg U/kg/day as uranyl fluoride. For the other uranium compounds studied, the maximum daily intakes that did not affect longevity were 1,130 mg U/kg/day as uranyl nitrate, 1,390 mg U/kg/day as uranium tetrafluoride, and 1,630 mg U/kg/day as uranium dioxide. About 18% of the experimental rats survived for the entire 2-year duration of the study, while about 38% of the control animals survived (Maynard and Hodge 1949). Most of the deaths in the available animal studies resulted from chemically induced renal damage.
The LD50 values for each species and other LOAEL values for mortality from exposure to uranium by the oral route are presented in Table 2-3 and plotted in Figure 2-3.
2.2.2.2 Systemic Effects
No human studies were located regarding respiratory, endocrine, dermal, ocular, body weight, or other systemic effects in humans following acute-, intermediate-, or chronic-duration oral exposure to uranium compounds.
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Animal data are lacking regarding musculoskeletal, metabolic, dermal, or ocular effects following oral exposure to uranium and its compounds for all durations. Similarly, no animal studies were located on the hematological effects of uranium in animals following acute-duration oral exposure or on the cardiovascular, endocrine, or other systemic effects following acute- or chronic-duration oral exposure. Data exist for the respiratory, renal, and body weight effects following oral exposure of animals to uranium for all durations. However, the existing data on the hematological, cardiovascular, hepatic, and other systemic effects of uranium in animals are limited to acute- or chronic-duration inhalation exposure; data on the gastrointestinal effects are limited to acute-duration exposure.
The highest NOAEL values and all reliable LOAEL values in each species and duration category for systemic effects from chemical exposures to uranium by the oral route are presented in Table 2-3 and plotted in Figure 2-3.
Respiratory Effects. Respiratory effects from oral exposure to uranium are unlikely. In an acute-duration animal study, no adverse effects on the respiratory system were reported in rats given single oral doses of 118 mg uranium per kilogram body weight per day (U/kg/day) as uranyl acetate dihydrate (Domingo et al. 1987).
In intermediate-duration exposures, Sprague-Dawley rats (10/sex/dose) were exposed to uranium as uranyl nitrate in the drinking water (males: up to 35.3 mg/kg/day; females: up to 40.0 mg/kg/day) for 28 days and then sacrificed. No treatment-related histopathological changes were found, and no changes in lung weights were noted (Gilman et al. 1998a). In several 30-day dietary studies, no adverse effects on the respiratory system were reported in rats exposed to 6,637 mg U/kg/day as uranyl nitrate hexahydrate, 8,769 mg U/kg/day as uranium tetrachloride, 11,033 mg U/kg/day as uranium peroxide, 10,611 mg U/kg/day as uranium tetrafluoride, 10,818 mg U/kg/day as uranyl fluoride, 12,342 mg U/kg/day as uranium dioxide, 8.167 mg U/kg/day as uranyl acetate dihydrate, or 11,650 mg U/kg/day as uranium trioxide (Maynard and Hodge 1949; Maynard et al. 1953; Stokinger et al. 1953). Lengthening the duration of exposure to uranium failed to produce detectable lesions in lungs of laboratory animals. Sprague-Dawley rats (15/sex/dose) were exposed to uranium as uranyl nitrate in drinking water (males: up to 36.73 mg/kg/day; females: up to 53.56 mg/kg/day) for 91 days and were sacrificed. No treatment-related histopathological changes were found in the lungs, and no changes in lung weights were noted (Gilman et al. 1998a). In addition, New Zealand rabbits were exposed to uranium as uranyl nitrate in the
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drinking water (males: up to 28.70 mg/kg/day; females: up to 43.02 mg/kg/day) for 91 days. No treatment-related histopathological changes were found, and no changes in lung weights were noted (Gilman et al. 1998b). Male New Zealand rabbits were also exposed to uranium as uranyl nitrate in drinking water (1.36 and 40.98 mg/kg/day) for 91 days, again with no histopathological or organ weight changes found (Gilman et al. 1998c).
In chronic-duration feeding studies, no adverse effects on the respiratory system were reported in 1-year studies of dogs given oral doses of 31 mg U/kg/day as uranium tetrachloride, 3,790 mg U/kg/day as uranium hexachloride, 8 mg U/kg/day as uranyl fluoride, or 4,407 mg U/kg/day as uranium dioxide (Maynard and Hodge 1949; Maynard et al. 1953). In 2-year studies, the respiratory system was unaffected in dogs and rats given 2 mg U/kg/day as uranyl nitrate hexahydrate and in rats given 12,141 mg U/kg/day as uranium dioxide, 664 mg U/kg/day as uranyl nitrate hexahydrate, 10,611 mg U/kg/day as uranium tetrafluoride, or 405 mg U/kg/day as uranyl fluoride (Maynard and Hodge 1949; Maynard et al. 1953; Stokinger et al. 1953).
Cardiovascular Effects. Cardiovascular effects following intake of uranium are unlikely. One case report documented a cardiovascular effect that was possibly related to uranium exposure in a male admitted to the hospital following deliberate ingestion of 15 g of uranyl acetate, along with an unknown quantity of benzodiazepine, in a failed suicide attempt. While body weight was not reported, the dose would be approximately 131 mg U/kg for a 70 kg reference man. Initial blood chemistry was unremarkable, and decreased cardiac output was consistent with ingestion of benzodiazepam. The patient was reported to have suffered from myocarditis resulting from the uranium ingestion, resolving 6 months after the ingestion (Pavlakis et al. 1996).
The available studies in animals have found no adverse cardiovascular effects following oral exposures for up to 30 days to uranium compounds. In one study, Sprague-Dawley rats (10/sex/dose) were exposed to uranium as uranyl nitrate in drinking water (males: up to 35.3 mg/kg/day; females: up to 40.0 mg/kg/day) for 28 days and sacrificed. No cardiac histopathological changes were found, and no changes in heart weights were noted (Gilman et al. 1998a). No changes in the heart or blood vessels were observed in rats following oral exposure to doses as high as 9,393 mg U/kg/day as uranyl nitrate hexahydrate, 8,769 mg U/kg/day as uranium tetrachloride, 11,033 mg U/kg/day as uranium peroxide, 10,611 mg U/kg/day as uranium tetrafluoride, 10,819 mg U/kg/day as uranyl fluoride,
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12,342 mg U/kg/day as uranium dioxide, 11,650 mg U/kg/day as uranium trioxide, or 7,859 mg U/kg/day as uranyl acetate dihydrate (Maynard and Hodge 1949). Sprague-Dawley rats (15/sex/dose) were exposed to uranium as uranyl nitrate in drinking water (males: up to 36.73 mg/kg/day; females: up to
53.56 mg/kg/day) for 91 days and sacrificed. No uranium-related histopathological changes were found in the heart, and no changes in heart weights were noted (Gilman et al. 1998a). In addition, New Zealand rabbits were exposed to uranium as uranyl nitrate in the drinking water (males: up to 28.70 mg/kg/day; females: up to 43.02 mg/kg/day) for 91 days. No treatment-related cardiac histopathological changes were noted, and no changes in heart weights were detected (Gilman et al. 1998b). Male New Zealand rabbits also were exposed to uranium as uranyl nitrate in drinking water (1.36 and 40.98 mg/kg/day) for 91 days, with no histopathological or organ weight changes (Gilman et al. 1998c).
Gastrointestinal Effects. A volunteer given a single dose of 1 g uranyl nitrate (14.3 mg/kg) and observed for clinical signs and symptoms within 24 hours after intake suffered acute nausea, vomiting, and diarrhea within a few hours of administration. All clinical signs returned to normal within 24 hours after administration of the oral uranyl nitrate dose (Butterworth 1955). Paralytic ileus was reported in a male after the deliberate ingestion of 15 g uranyl acetate (Pavlakis et al. 1996). While body weight was not reported, the dose would be approximately 131 mg U/kg for a 70 kg reference man. No other reports of gastrointestinal effects after acute-duration exposure to uranium in either humans or laboratory animals were available.
Studies of intermediate-duration exposure to uranium compounds were available for laboratory animals. In one study, Sprague-Dawley rats (10/sex/dose) were exposed to uranium as uranyl nitrate in the drinking water (males: up to 35.3 mg/kg/day; females: up to 40.0 mg/kg/day) for 28 days and sacrificed. No treatment-related histopathological changes were found, and no changes in organ weights were noted (Gilman et al. 1998a). In a study of longer duration, Sprague-Dawley rats (15/sex/dose) were exposed to uranium as uranyl nitrate in drinking water (males: up to 36.73 mg/kg/day; females: up to
53.56 mg/kg/day) for 91 days and then sacrificed. No treatment-related histopathological changes were found in the gastrointestinal tract, and no changes in stomach and intestinal weights were noted (Gilman et al. 1998a). In addition, New Zealand rabbits were exposed to uranium as uranyl nitrate in the drinking water (males: up to 28.70 mg/kg/day; females: 0 up to 43.02 mg/kg/day) for 91 days. No treatment-related histopathological changes were found, and no changes in organ weights (i.e., colon, duodenum, stomach [gastric cardia, fundus, and pylorus]) were noted (Gilman et al. 1998b). Male New Zealand
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rabbits were exposed to uranium as uranyl nitrate in drinking water (1.36 and 40.98 mg/kg/day) for 91 days, with no histopathological or organ weight changes found (Gilman et al. 1998c).
Hematological Effects. In one case report, a male (no age or weight given), was admitted to hospital following the deliberate ingestion of 15 g of uranyl acetate, along with an unknown quantity of benzodiazepine, in a failed suicide attempt. While body weight was not reported, the dose would be approximately 131 mg U/kg for a 70 kg reference man. Initial blood chemistry was unremarkable; however, an anemia developed and continued to progress over the next 8 weeks, along with persistent renal dysfunction (Pavlakis et al. 1996). While the authors attributed the renal dysfunction to uranium ingestion, the etiology of the anemia was unknown. The patient also suffered from peptic ulcer disease which may have been related to the anemia.
The majority of animal studies show no effect of uranium on hematological parameters after oral exposure. Exposure to uranium as uranyl nitrate in drinking water had no hematological effects in Sprague-Dawley rats after 28 days (up to 40 mg U/kg/day) or 91 days (up to 53 mg U/kg/day) (Gilman et al. 1998a), or in New Zealand rabbits after 91 days (up to 43 mg U/kg/day) (Gilman et al. 1998b, 1998c). Exposure to a variety of uranium compounds in feed had no effect on hematological parameters in intermediate- and chronic-duration studies (Maynard and Hodge 1949). One study reported a significant increase in the hematocrit and hemoglobin values, the mean corpuscular hemoglobin concentration, and the number of erythrocytes at 9 mg U/kg/day as uranyl acetate in drinking water for 4 weeks, but not at
4.5 mg U/kg/day and lower doses (Ortega et al. 1989a).
In a chronic-exposure feeding study, mild anemia and an increased leucocyte count were observed in rats given uranyl nitrate hexahydrate doses corresponding to 33 mg U/kg/day for 2 years (Maynard and Hodge 1949; Maynard et al. 1953).
Musculoskeletal Effects. In one human case report, a human male (no age or weight given), was admitted to hospital following the deliberate ingestion of 15 g of uranyl acetate, along with an unknown quantity of benzodiazepine, in a failed suicide attempt. While body weight was not reported, the dose would be approximately 131 mg U/kg for a 70 kg reference man. The patient suffered from increasing rhabdomyolysis (biochemically characterized by increased creatine kinase). At 6 months following the
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initial toxic insult, the rhabdomyolysis had resolved, and the subject showed no residual signs of muscle toxicity (Pavlakis et al. 1996). The etiology of this effect is unknown.
In the available animal studies, the existing data provide evidence that uranium exposure does not cause detectable damage to the musculoskeletal system. Histopathological examination of muscle after exposure to uranium in drinking water as uranyl nitrate showed no effects in Sprague-Dawley rats after 28 days (up to 40 mg U/kg/day) or 91 days (up to 53 mg U/kg/day) (Gilman et al. 1998a), or in New Zealand rabbits after 91 days (up to 43 mg U/kg/day) (Gilman et al. 1998b, 1998c).
Hepatic Effects. Few human data are available on the hepatic effects of uranium. In one case report, a human male (no age or weight given), was admitted to hospital following the deliberate ingestion of 15 g of uranyl acetate, along with an unknown quantity of benzodiazepine, in a failed suicide attempt. While body weight was not reported, the dose would be approximately 131 mg U/kg for a 70 kg reference man. The patient suffered from increasing liver dysfunction, characterized by increased serum levels of ALT, AST, and GGK. Six months following the initial toxic insult, the patient had no residual signs of hepatic toxicity (Pavlakis et al. 1996). The etiology of this effect is unknown, although histological signs of hepatic toxicity have been observed in animals after oral exposure to uranium.
In the available animal studies, the existing data provide evidence that uranium exposure can damage the liver, although the etiology for this effect is not certain. In an acute-duration study in which Sprague-Dawley rats were given single gavage doses of 5.6 or 118 U/kg as uranyl acetate dihydrate, microhemorrhagic foci in the liver were observed at both doses tested (Domingo et al. 1987).
Ingested uranium doses were also hepatotoxic to dogs in studies of intermediate-duration exposure. When uranyl fluoride was tested at 7.7, 15.4, 77.3, 386.7, or 3,864 mg U/kg/day for 30 days, fatty infiltration was seen in dogs at the 15.4 mg U/kg/day dose level (Maynard and Hodge 1949). In other tests, uranium tetrachloride induced minimal hepatic lesions at a dose level of 313 mg U/kg/day; uranium peroxide induced mild degeneration at a dose level of 386 mg U/kg/day; uranium dioxide induced mild degeneration at a dose level of 441 mg U/kg/day; uranium trioxide induced slight fatty infiltration at a dose level of 416 mg U/kg/day; triuranium octaoxide induced mild fatty changes at a dose level of 5,653 mg U/kg/day; sodium diuranate induced mild degeneration at a dose level of 37.5 mg U/kg/day;
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uranium tetrafluoride caused degenerative fatty changes at a dose level of 15,159 mg U/kg/day; and uranyl nitrate hexahydrate induced minimal hepatic degeneration at a dose level of 237 mg U/kg/day (Maynard and Hodge 1949).
Hepatic toxicity was also found several other studies. In one study, Sprague-Dawley rats (15/sex/dose) were exposed to uranium as uranyl nitrate in drinking water (males: up to 36.73 mg/kg/day; females: up to 53.56 mg/kg/day) for 91 days and then sacrificed. Hepatic lesions, which included anisokaryosis, vesiculation, increased portal density, perivenous vacuolation, and homogeneity, were observed in the liver at all doses (Gilman et al. 1998a), although the dose ranging portion of this study found no effects at essentially the same doses as those discussed below (Gilman et al. 1998c). However, in New Zealand rabbits exposed to uranium as uranyl nitrate in the drinking water (males: 0, 0.05, 0.20, 0.88, 4.82, and
28.70 mg/kg/day; females: 0, 0.49, 1.32, and 43.02 mg/kg/day) for 91 days, no treatment-related histopathological changes were found, and no changes in liver weights were noted (Gilman et al. 1998b). In contrast, another study by the same investigator in male New Zealand rabbits exposed to uranium as uranyl nitrate in drinking water (1.36 and 40.98 mg/kg/day) for 91 days found irregular accentuation of zonation in the liver, accompanied by increased variation in hepatocellular nuclear size, nuclear pyknosis, and extensive cytoplasmic vacuolization. These changes were found to be treatment-related but not dose-related (Gilman et al. 1998c).
In other intermediate-duration studies, no effects were seen on the liver of dogs given oral doses of 9,393 mg U/kg/day as uranyl nitrate hexahydrate or 191 mg U/kg/day as ammonium diuranate for 30 days (Maynard and Hodge 1949). Similarly, no effects were seen on the liver of rats given oral doses of 8,769 mg U/kg/day as uranium tetrachloride, 11,033 mg U/kg/day as triuranium peroxide, 10,818 mg U/kg/day as uranyl fluoride, 12,342 mg U/kg/day as uranium dioxide, 11,650 mg U/kg/day as triuranium trioxide, or 7,859 mg U/kg/day as uranium acetate dihydrate for 30 days (Maynard and Hodge 1949). Sprague-Dawley rats (10/sex/dose, 60 g) were exposed to uranium as uranyl nitrate in drinking water (males: up to 35.3 mg/kg/day; females: up to 40.0 mg/kg/day) for 28 days and then sacrificed. No treatment-related histopathological changes were found, and no changes in liver weights were noted (Gilman et al. 1998a).
Renal Effects. Uranium has been identified as a nephrotoxic metal, exerting its toxic effect by chemical action mostly in the renal proximal tubules of humans and animals. In this regard, uranium is a less potent nephrotoxin than the classical nephrotoxic metals (cadmium, lead, mercury) (Goodman 1985).
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Few human data are available that adequately describe the dose-response toxicity of uranium after an oral exposure. In one human case report study, a male (no age or weight given), was admitted to hospital following the deliberate ingestion of 15 g of uranyl acetate, along with an unknown quantity of benzodiazepine, in failed a suicide attempt. While body weight was not reported, the dose would be approximately 131 mg U/kg for a 70 kg reference man. Initial blood chemistry was normal; however, 16 hours after admission, his blood urea levels had doubled and creatinine levels had increased 3.5-fold, which suggested renal damage. A diagnosis of acute nephrotoxicity from heavy metal exposure was made, and chelation therapy with Ca-EDTA, sodium bicarbonate, and mannitol was initiated. His plasma uranium on the day following commencement of chelation therapy was 3.24 μmol/L, decreasing to 1.18 μmol/L after 5 days of chelation and dialysis. Chelation therapy was then stopped; however, dialysis continued for 2 weeks at which point kidney function recovered sufficiently to allow withdrawal of dialysis therapy. The patient’s anemia persisted over the next 8 weeks, along with persistent renal dysfunction. Additional chelation therapy was initiated with both Ca EDTA and Ca DTPA (diethylenetriamine pentaacetic acid) without success. At 6 months following the initial toxic insult, the patient still suffered from an incomplete Fanconi syndrome (renal tubular acidosis) requiring supplemental sodium bicarbonate therapy on a daily basis (Pavlakis et al. 1996). The authors suggested that pre-existing peptic ulcer disease in this patient may have exacerbated toxicity by increased absorption of uranium through the damaged stomach mucosal layer.
Although there is little additional information about renal effects in humans following oral exposure to uranium compounds, there is sufficient information in animals with high exposures to both soluble and insoluble uranium to permit the conclusion that uranium has a low order of metallotoxicity in mammals. Many of the nonradioactive heavy metals such as lead, arsenic, and mercury would produce severe, perhaps fatal, injury at the levels of exposure reported for uranium in the literature. The negative findings regarding renal injury among workers exposed over long time periods to insoluble compounds are particularly significant in view of the high levels of exposure reported (Eisenbud and Quigley 1955). The pathogenesis of the kidney damage in animals indicates that regeneration of the tubular epithelium occurs in survivors upon discontinuation of exposure to uranium (Bentley et al. 1985; Dygert 1949b; Maynard and Hodge 1949; Pozzani 1949; Rothermel 1949; Rothstein 1949c; Spiegl 1949; Stokinger et al. 1953).
Male Sprague-Dawley rats exposed to a single gavage dose of 5.6 mg U/kg suffered slight renal dysfunction and minimal microscopic lesions in the tubular epithelium (Domingo et al. 1987).
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An intermediate-duration oral study in which dogs were given doses of 37.5 or 187 mg U/kg/day as sodium diuranate in the diet for 30 days found elevated peak NPN and BUN but not in a dose-dependent manner. Blood sugar was also slightly elevated. Necropsy findings revealed mild degeneration and necrosis in the kidneys at the higher dose level but only minimal degeneration and necrosis at
37.5 mg U/kg/day (Maynard and Hodge 1949). In other animal studies, exposure to uranium (uranyl fluoride, triuranium octaoxide, uranyl nitrate hexahydrate, uranium tetrachloride, uranium peroxide, ammonium diuranate) at oral doses as low as 0.05 mg U/kg/day and as high as 7,859 mg U/kg/day for 30 days were damaging to the kidneys. Nephrotoxic effects found in these animals ranged from minimal microscopic lesions in the tubular epithelium (for low doses) to extensive necrosis in the tubular epithelium (for high doses of soluble compounds) (Maynard and Hodge 1949). No effects on the kidneys were found in rats similarly exposed to 12,342 mg U/kg/day as uranium dioxide or 11,650 mg U/kg/day as uranium trioxide for 30 days (Maynard and Hodge 1949); perhaps, this finding was due to the low gastrointestinal absorption of the insoluble salt.
In intermediate-duration studies, dogs orally exposed to up to 95 mg U/kg/day as uranyl nitrate hexahydrate for 138 days suffered elevated NPN, BUN, glucosuria, and proteinuria at doses of 95 mg U/kg/day and higher, no effect was seen at 47 mg U/kg/day (Maynard and Hodge 1949). Exposure of mice to 452 mg U/kg/day as uranyl fluoride for 48 weeks resulted in the kidneys being tan-gray in color with nodules on the surface (Tannenbaum and Silverstone 1951). However, the kidneys appeared to be normal upon microscopic examination. In other studies, Sprague-Dawley rats (10/sex/dose) were exposed to uranium as uranyl nitrate in drinking water (males: 0.05, 0.27, 1.34, 6.65, 35.3 mg U/kg/day; females: 0.07, 0.33, 1.65, 7.82, 40.0 mg U/kg/day) for 28 days and then sacrificed. No treatment-related histopathological changes were found, and no changes in organ weights were noted. The only effect observed was a significant increase in serum uric acid in females at 40 mg U/kg/day (1.64 vs. 1.18 mg/dL in controls). This 28-day dose range finding study found few adverse effects at even the highest dose, but was followed by a 91-day study of the same regimen, with significantly different results. In that study, Sprague-Dawley rats (15/sex/dose) exposed to uranium as uranyl nitrate in drinking water (males: <0.0001, 0.06, 0.31, 1.52, 7.54, 36.73 mg U/kg/day; females: <0.0001, 0.09, 0.42, 2.01, 9.98, 53.56 mg U/kg/day) for 91 days were found to have renal lesions of the tubules (apical nuclear displacement and vesiculation, cytoplasmic vacuolation, and dilation), glomeruli (capsular sclerosis), and interstitium (reticulin sclerosis and lymphoid cuffing) observed in the lowest exposure groups. No explanation for the differences was provided (Gilman et al. 1998a).
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Two studies by MacDonald-Taylor et al. (1992, 1997) produced similar renal lesions in rabbits. In these studies, weanling New Zealand male rabbits were exposed to uranium for 91 days via drinking water containing 0, 24, or 600 mg/L uranyl nitrate Doses were not calculated from water intake. Calculations using default reference values for this species result in doses of 0, 0.93, and 23 mg U/kg/day (EPA 1998). Each treatment group was divided into 3 subgroups: immediate sacrifice and either 45-day or 91-day recovery period. At the end of the recovery periods, rabbits were sacrificed and renal sections prepared for electron microscopy. Thickness of the glomerular basement membrane (GBM) was measured from electron micrographs. Measurements were taken at approximately 35 μm increments; 600–900 measurements were made for each treatment group and recovery period. Uranyl nitrate exposure resulted in thickening of the membrane in the rabbits. Control thickness was approximately 80 μm; the thickness was 96.3 μm immediately after exposure at the low dose and increased to 103 μm after a 91-day recovery. Initial thickness after 91 days exposure in the high-dose group was 109 μm and had increased to 117 μm after a 91-day recovery period. Similarly, New Zealand rabbits were exposed to uranium as uranyl nitrate in the drinking water (males: up to 28.70 mg U/kg/day; females: up to 43.02 mg U/kg/day) for 91 days (Gilman et al. 1998b). Dose-dependent differences consisted of histopathological changes limited primarily to kidney and were more pronounced in male rabbits. In the males, a significant increased incidence of anisokaryosis and nuclear vesiculation was observed in all treated groups. Nuclear pyknosis and hyperchromicity were observed in all treated groups except in the 0.05 mg U/kg/day treatment group. Tubular dilation, atrophy, protein casts, and collagen sclerosis were observed at 4.82 and 28.70 mg U/kg/day. Reticulin sclerosis was observed at 0.88, 4.82, and 28.70 mg U/kg/day. Anisokaryosis and nuclear vesiculation were observed in all treated female groups. Tubular dilation and atrophy were also observed. Collagen sclerosis was observed at 43.02 mg U/kg/day, reticulin sclerosis was observed at 0.49 and 43.02 mg U/kg/day. Females drank 65% more water than the males; however, the females appeared to be less affected by the exposure regimen. These exposed females did develop significant tubular nuclear pathology in the lowest exposure group, but not to the degree of the exposed males (Gilman et al. 1998b). The LOAEL of 0.5 mg U/kg/day from this study was used to develop an intermediate-duration MRL of 2.0×10-3 mg/kg/day for oral exposure to uranium and is shown in Table 2-3 and plotted in Figure 2-3.
In another study, male New Zealand rabbits were exposed to uranium as uranyl nitrate in drinking water
(1.36 and 40.98 mg U/kg/day) for 91 days, and were then allowed to recover for several weeks after dosing ceased (Gilman et al. 1998c). No differences in urinary parameters were noted in any of the groups exposed to the 1.36 mg U/kg/day dose. Kidney weight as a percentage of body weight was
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significantly increased in the 40.98 mg U/kg/day group (compared to controls) immediately after exposure, but was not significantly increased after 45 days of exposure. In rabbits exposed to the 40.98 mg U/kg/day dose, urinary output was decreased at week 1, with increased excretion of glucose, protein and leucine aminopeptidase activity. Similar results were observed at week 4 after dosing began. Seven days after the start of the recovery period, urinary volume was increased and glucose secretion remained elevated. Protein and leucine aminopeptidase activity excretion returned to normal. At 3, 5, and 13 weeks post-exposure, urinary parameters were normal. Groups exposed to 40.98 mg U/kg/day had increases in percentage and total lymphocyte counts after the 91-day recovery period but not at the end of exposure. Focal dilation of renal proximal tubules was observed in both treated groups accompanied by cytoplasmic vacuolation. Nuclear changes included apical displacement and irregular placement with vesiculation, anisokaryosis, and pyknosis. Tubular basement membranes were normal early in injury but thickened focally during recovery. Changes induced by exposure at 40.98 mg U/kg/day persisted for up to 45 days and in some cases for 91 days (Gilman et al. 1998c).
Endocrine Effects. No endocrine effects after oral intake of uranium have been reported in humans. Few endocrine effects have been reported after uranium exposure in laboratory animals. In animal studies, a dose of 0.07 mg U/kg/day as uranyl nitrate hexahydrate for 16 weeks in drinking water resulted in degenerative changes in the thyroid epithelium and altered thyroid function in Wistar rats (Malenchenko et al. 1978). Sprague-Dawley rats (10/sex/dose) were exposed to uranium as uranyl nitrate in drinking water (males: up to 35.3 mg U/kg/day; females: up to 40.0 mg U/kg/day) for 28 days and then sacrificed. No treatment-related histopathological changes were found in any of the endocrine organs studied (adrenal, pancreas, parathyroid, pituitary, thymus, thyroid), and no treatment-related changes in these organ weights were noted (Gilman et al. 1998a). In addition, New Zealand rabbits were exposed to uranium as uranyl nitrate in the drinking water (males: up to 28.70 mg U/kg/day; females: up to 43.02 mg U/kg/day) for 91 days. No treatment-related histopathological changes were found, and no weight changes in the adrenal, pancreas, parathyroid and pituitary glands were noted (Gilman et al. 1998b). Male New Zealand rabbits exposed to uranium as uranyl nitrate in drinking water (1.36 and 40.98 mg U/kg/day) for 91 days also failed to show any treatment-related histopathological or organ weight changes (Gilman et al. 1998c). In another study, Sprague-Dawley rats (15/sex/dose) were exposed to uranium as uranyl nitrate in the drinking water (males: up to 36.73 mg U/kg/day; females: up to 53.56 mg U/kg/day) for 91 days. Thyroid lesions were observed in both sexes (multifocal reduction of follicular
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size, increased epithelial height in males at 0.31 mg U/kg/day and females at 2.01 mg U/kg/day). A decreased amount and density of colloid in the thyroid was observed in males only.
Body Weight Effects. No body weight effects after oral intake of uranium have been reported in humans.
Oral exposure to uranium compounds has caused body weight effects in animals, but these effects are not necessarily the result of systemic toxicity. The initial loss of body weight observed in animals exposed to high doses of uranium in the diet in acute-, intermediate-, and chronic-duration studies is usually accompanied by decreased food consumption in these animals. The decreased food consumption could be due to the aversive taste of uranium compounds in food. Subsequent acclimatization of the animals to the taste would normalize food intake and, consequently, reverse the initial loss of body weight. Thus, the changes in body weight seen in such studies may be due more to reduction in food consumption due to distaste than to uranium-specific chemical or radiological toxicity. More recent studies in which sugar was added to the drinking water of animals to remove the aversive effect of uranium (Ortega et al. 1989a) support this hypothesis.
Rats given single oral doses of 664 mg U/kg as uranyl nitrate hexahydrate or 55 mg U/kg as uranium peroxide (Maynard et al. 1953), 7,859 mg U/kg as uranium acetate dihydrate for 30 days (Maynard and Hodge 1949), or 664 mg U/kg as uranyl nitrate hexahydrate for 30 days in the feed suffered unspecified decreases in body weight gain (Maynard et al. 1953). Similarly, body weight losses of 18, 35, 27, 20, and 29%, respectively, were observed in rats given oral doses of 886 mg U/kg/day as uranium tetrachloride, 1,081 mg U/kg/day as uranyl fluoride, or 664 mg U/kg/day as uranyl nitrate hexahydrate for 30 days (Maynard and Hodge 1949); rabbits given oral doses of 14.2 mg U/kg/day as uranyl nitrate hexahydrate for 30 days (Maynard and Hodge 1949); and rats given oral doses of 270 mg U/kg/day as uranyl fluoride for 2 years (Maynard and Hodge 1949; Maynard et al. 1953).
No harmful effects on body weight were seen in rats given 12,342 mg U/kg as uranium dioxide or 11,650 mg U/kg as uranium trioxide for 30 days (Maynard and Hodge 1949), mice given 1,100 mg U/kg as uranyl nitrate hexahydrate for 18 weeks or 462 mg U/kg as uranyl nitrate hexahydrate for 48 weeks (Tannenbaum and Silverstone 1951), or in Sprague-Dawley rats exposed to uranium as uranyl nitrate in drinking water at doses up to 35.3 mg U/kg/day (males) and 40 mg U/kg/day (females) for 28 days or up
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to 36.73 mg U/kg/day (males) and 53.56 mg U/kg/day (females) for 91 days (Gilman et al. 1998a). No alterations in body weights were observed in rats given 12,341 mg U/kg as uranium dioxide or 10,611 mg U/kg as uranium hexafluoride for 2 years, or dogs given 8 mg U/kg as uranyl fluoride or 95 mg U/kg as uranyl nitrate hexahydrate for 1 year (Maynard and Hodge 1949; Maynard et al. 1953). In animal studies, reduced food intake was observed following a single oral dose of 5.6 mg U/kg as uranyl nitrate hexahydrate to rats (Domingo et al. 1987) and in a 48-week study in rats and mice at 1,100 mg U/kg/day as uranyl nitrate hexahydrate (Tannenbaum and Silverstone 1951). It has been suggested that this reduced food intake is a result of loss of appetite due to the unpalatability of the uranium compounds in the animals' food (Dygert 1949e).
2.2.2.3 Immunological and Lymphoreticular Effects
No information was located regarding the effects of uranium on the immune system in humans following oral exposure for any duration.
In laboratory animals, oral exposure of rats, mice, and rabbits to uranium had no significant effect on immune system function. In one study Sprague-Dawley rats (10/sex/dose) were exposed to uranium as uranyl nitrate in drinking water (males: up to 35.3 mg U/kg/day; females: up to 40.0 mg U/kg/day) for 28 days and then sacrificed. No treatment-related effects were noted in the immunological/lymphoreticular tissues examined (bone marrow, mesenteric and mediastinal lymph nodes, spleen, and thymus) (Gilman et al. 1998a). In addition, New Zealand rabbits were exposed to uranium as uranyl nitrate in the drinking water (males: up to 28.70 mg U/kg/day; females: up to 43.02 mg U/kg/day) for 91 days. No histopathological changes were found, and no changes in the bone marrow, mesenteric and mediastinal lymph nodes, or thymus were noted (Gilman et al. 1998b). Rats exposed to oral doses of 0.07 mg U/kg as uranyl nitrate hexahydrate for 4 weeks showed an increase in spleen weight but the body weights of both the control and test animals were not provided, making it impossible to determine whether the net change in spleen weight had any toxicological significance (Malenchenko et al. 1978). Sprague-Dawley rats exposed to uranium as uranyl nitrate in drinking water (males: up to 36.73 mg U/kg/day; females: up to 53.56 mg U/kg/day) for 91 days showed sinus hyperplasia of the spleen in both sexes at the highest dose (males: 36.73; females: 53.56 mg U/kg/day). No lesions were observed in bone marrow, mesenteric and medistinal lymph nodes, or in the thymus (Gilman et al. 1998a). In other studies with mice and rats, no histological changes in the spleen, lymph nodes, or bone marrow were seen
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in the animals following administration of up to 5,000 mg U/kg of various uranium compounds (uranyl nitrate hexahydrate, uranyl fluoride, uranium dioxide, uranium peroxide, uranium tetrafluoride, uranium tetrachloride, triuranium octaoxide, or uranium trioxide) in the diet for 48 weeks or 2 years. No consistent hematological changes were found in hematocrit, hemoglobin, or white blood cell counts (Maynard et al. 1953; Tannenbaum and Silverstone 1951). No other specific immunological tests were performed.
2.2.2.4 Neurological Effects
No studies were located for humans regarding neurological effects following oral exposure to uranium compounds.
No evidence of histological effects in nervous tissue have been reported after oral exposure to uranium compounds in animal studies, although one study reported clinical signs of neurotoxicity. Piloerection, tremors, hypothermia, pupillary size decreases and exophthalmos were seen at all dose levels in a study with Sprague-Dawley rats given single gavage doses of 11, 22, 45, 90, 179, 358, or 717 mg U/kg as uranyl acetate dihydrate. The signs became more severe as the number of days post-treatment increased (Domingo et al. 1987). In another study, Sprague-Dawley rats (10/sex/dose) were exposed to uranium as uranyl nitrate in drinking water (males: up to 35.3 mg U/kg/day; females: up to 40.0 mg U/kg/day) for 28 days and then sacrificed. No treatment-related effects were noted in the three sections of brain examined histopathologically (Gilman et al. 1998a). No treatment-related effects on the brains of Sprague-Dawley rats (15/sex/dose) exposed to uranium as uranyl nitrate in the drinking water (males: up to 36.73 mg U/kg/day; females: up to 53.56 mg U/kg/day) for 91 days were found (Gilman et al. 1998a). Additionally, New Zealand rabbits exposed to uranium as uranyl nitrate in the drinking water (males: up to 28.70 mg U/kg/day; females: up to 43.02 mg U/kg/day) for 91 days showed no brain histopathological changes.
The LOAEL value for this study is presented in Table 2-3 and plotted in Figure 2-3.
2.2.2.5 Reproductive Effects
No human studies were located regarding reproductive effects following oral exposure to uranium compounds. Limited animal studies have shown some effects on reproductive function but generally no evidence of histopathological damage to reproductive tissues. No reproductive effects or changes in
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reproductive organ weights were found in the epididymis, testes, ovary, or uterus of Sprague-Dawley rats (10/sex/dose) exposed to uranium as uranyl nitrate in the drinking water (males: up to 35.3 mg U/kg/day; females: up to 40.0 mg U/kg/day) for 28 days (Gilman et al. 1998a). No reproductive effects or changes in reproductive organ weights were found in the epididymis, testes, ovary, or uterus of Sprague-Dawley rats (15/sex/dose) exposed to uranium as uranyl nitrate in drinking water (males: up to 36.73 mg U/kg/day; females: up to 53.56 mg U/kg/day) for 91 days (Gilman et al. 1998a). New Zealand rabbits exposed to uranium as uranyl nitrate in the drinking water (males: up to 28.70 mg U/kg/day; females: up to 43.02 mg U/kg/day) for 91 days showed no histopathological or organ weight changes in the epididymis, ovary, testes, or uterus (Gilman et al. 1998b). No effects on fertility were found in mice given oral gavage doses of 14 mg U/kg/day as uranyl acetate dihydrate in a 4- to 8-week study (Paternain et al. 1989). In a 64-day study with Swiss-Webster mice, no significant differences in the total implantations, early and late resorptions, or the number of live and dead fetuses were observed in females mated with male mice treated with drinking water doses of 45 mg U/kg/day as uranyl acetate dihydrate, as compared to untreated controls; but a reduced sperm count was observed in the 11.2 mg U/kg/day group (Llobet et al. 1991). However, in another study, offspring of male Swiss mice exposed to 2.8, 5.6, or 14 mg U/kg/day intragastrically as uranyl acetate dihydrate for 38–60 days before mating with female mice that had received the same doses orally for 14 days prior to mating exhibited reproductive abnormalities manifested as reduced implantations and increased fetal resorptions. The average number of total implantations was only different in the 2.8 mg U/kg/day group (Paternain et al. 1989). Maternal toxicity (reduced weight gain and food consumption, increased relative liver weight) was seen at all doses in 20 pregnant Swiss mice given uranyl acetate dihydrate (3, 6, 14, or 28 mg U/kg/day) by gavage on gestation days (Gds) 6–15 and sacrificed on Gd 18 to assess potential maternal and fetal toxicity (Domingo et al. 1989a).
In chronic-duration studies, male rats given high oral doses (331 mg U/kg/day) of uranyl nitrate hexahydrate in the diet for 2 years developed testicular degeneration; female rats given oral doses of 664 mg U/kg/day as uranyl nitrate hexahydrate for 2 years had reduced litter sizes (Maynard et al. 1953). Since incidence and dose-response data were not provided in this report, its significance is unclear.
The highest NOAEL values and all reliable LOAEL values in each species and duration category for reproductive effects from exposure to uranium by the oral route are presented in Table 2-3 and plotted in Figure 2-3.
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2.2.2.6 Developmental Effects
No studies were located that reported developmental effects in humans following oral exposure to uranium for any duration. Animal studies indicate that oral exposure to uranium can cause developmental effects, but only at relatively high doses.
In animal studies, pregnant Swiss mice were exposed to uranium as uranyl acetate dihydrate by gavage in water at a dose of 0.028, 0.28, 2.8, 28 mg U/kg/day from day 13 of gestation through postnatal day 21. Treatment had no significant effects on mean litter size at birth or on day 4, but litter size was significantly decreased at postnatal day 21 at 28 mg U/kg/day (5.5 vs. 8.8 in water-only controls). The viability index (number of pups viable at day 21/number of pups born) and the lactation index (number of pups viable at day 21/number of pups retained at day 4) were significantly decreased in the 28 mg U/kg/day group. No significant differences were observed in developmental signs (pinnae unfolding, lower incisor eruption, eye opening), or in pup weight or body length (Domingo et al. 1989b). Structural variations were not assessed in this report.
The offspring of male Swiss mice exposed to 2.8, 5.6, and 14 mg U/kg/day intragastrically as uranyl acetate dihydrate for 38–60 days before mating with female mice that had received the same doses orally for 14 days prior to mating exhibited developmental defects. The average number of total implantations was only different in the 2.8 mg U/kg/day group. The numbers of late resorptions and dead fetuses were significantly increased for the 14 mg U/kg/day group. Significantly reduced viability was observed in the
5.6 mg U/kg/day group. A dose-response relationship was observed for reduced offspring growth as determined by body weight and body length (Paternain et al. 1989). Similarly, a dose-related fetotoxicity, manifested as reduced fetal body weight and length, an increase in the incidence of stunted fetuses and external and skeletal malformations, and developmental variations, was reported in the offspring of 20 pregnant Swiss mice given uranyl acetate dihydrate (3, 6, 14, and 28 mg U/kg/day) by gavage on Gds 6–15 and sacrificed on Gd 18. External malformations included a significant increase in the incidence of cleft palate ($6 mg U/kg/day) and hematomas (at 3 and 28 mg U/kg/day). Underdeveloped renal papillae were seen in the 3 and 14 mg U/kg/day groups. An increase in the incidence of skeletal abnormalities (bipartite sternebrae and reduced or delayed ossification of the hind limb, fore limb, skull, and tail) were seen in the 14 and 28 mg U/kg/day groups. Embryolethality was not found at any of the dose levels tested (Domingo et al. 1989a); however, in another study, embryolethality was found in
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offspring of mice given an oral gavage dose of 14 mg U/kg/day as uranyl acetate dihydrate in a 4–8-week study (Paternain et al. 1989).
In another study, the mean litter size of the offspring of female Sprague-Dawley rats was significantly lower (p<0.05) at an oral exposure of 28 mg U/kg/day uranyl acetate dihydrate on postnatal day 21 when a group of rats were exposed to 0.028, 0.38, 2.8, or 28 mg U/kg/day for 30 days. The viability index (day 21:day 0) and lactation index were also significantly reduced at this exposure level. No differences in the developmental milestones monitored (pinna attachment, eye opening, incisor eruption) were observed in the treated animals. Treatment with uranium had no significant effect on length of gestation and sex ratios and on mean litter size at birth or postnatal day 4 as well as on body weight or pup body length throughout lactation. There was no significant effect on food consumption during the periods of late gestation and lactation (Domingo et al. 1989b).
The highest NOAEL values and all reliable LOAEL values in each species and duration category for developmental effects from exposure to uranium by the oral route are presented in Table 2-3 and plotted in Figure 2-3.
2.2.2.7 Genotoxic Effects
No information was located regarding the toxic action of uranium on genetic material in humans or animals following oral exposure for any duration.
Because uranium is a predominantly alpha-emitting radionuclide, current theories on gene mutation and chromosomal aberrations by high-LET alpha radiation suggest a potential for genotoxicity from uranium’s radioactivity (BEIR 1980, 1988, 1990; Leach et al. 1970; Morris et al. 1990; Muller et al. 1967; Otake and Schull 1984; Sanders 1986; Stokinger et al. 1953; UNSCEAR 1982, 1986, 1988) (see Appendix D for a review of the hazards associated with radionuclide exposure). Other genotoxicity studies are discussed in Section 2.5.
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2.2.2.8 Cancer
No evidence linking oral exposure to uranium to human cancer has been found. Although natural, depleted, or enriched uranium and uranium compounds have not been evaluated in rodent cancer bioassays by any route by the NTP (BEIR 1980, 1988, 1990; Hahn 1989; Sanders 1986; UNSCEAR 1982, 1986, 1988), there is potential for the carcinogenicity of uranium, since it emits primarily alpha radiation. Nevertheless, no evidence has been found to associate human exposure to uranium compounds and carcinogenesis. The National Academy of Sciences has determined that bone sarcoma is the most likely cancer from oral exposure to uranium; however, their report noted that this cancer has not been observed in exposed humans and concluded that exposure to natural uranium may have no measurable effect (BEIR IV).
Similarly, the results of several oral studies with uranium in several species were negative for evidence of cancer induction (Maynard and Hodge 1949; Maynard et al. 1953; Tannenbaum and Silverstone 1951).
No studies were located that provided evidence that oral exposure of humans to uranium as an alpha-emitting radiation source causes cancer. The available human data on the relative potential of ingested radium and uranium isotopes to induce cancers in humans concluded that the cumulative lifetime risk to 1 million people, each ingesting 5 pCi of a radium isotope (226Ra, 228Ra, and 224Ra) per day, for the induction of skeletal cancers (bone sarcomas and carcinomas of the head sinuses) is 9 bone sarcomas and 12 head carcinomas for 226Ra, 22 bone sarcomas for 228Ra, and 1.6 bone sarcomas for 224Ra. Assuming that the risk per rad of the average skeletal dose is equal for 226Ra and uranium isotopes with half-lives exceeding 1,000 years, and that the equilibrium skeletal content is 25 times the daily ingestion of 226Ra but 11 times the daily ingestion of long-lived uranium, the cumulative lifespan risk to 1 million people, each ingesting 5 pCi per day of 234U (0.0008 μg), 235U (2.3 μg), or 238U (15 μg), is estimated to be about
1.5 bone sarcomas. However, no cancers would be expected if the incidence is found to vary with the square of the dose instead of linearly (Mays et al. 1985). The BEIR IV report came to the same conclusion, but reserved the opinion that bone sarcomas might be caused by highly enriched uranium. The report estimated a lifetime risk of excess bone sarcomas per million people of 1.5 if soluble uranium isotopes were ingested at a constant daily rate of 1 pCi/day (0.037 Bq/day). The number of bone sarcomas that occur naturally in a population of a million people is 750. However, no quantitative risk coefficient estimates for developing human exposure protection benchmarks were provided in this report. In addition, the BEIR IV analysis was presumably based on generic short-lived alpha-emitting sources,
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such as radon that have a higher potential for inducing cancer, and not on radionuclides with relatively longer radioactive half-lives like 238U, 235U, and 234U. Perhaps more importantly, the BEIR IV report concluded that "…exposure to natural uranium is unlikely to be a significant health risk in the population and may well have no measurable effect" (BEIR IV 1988).
The available long-term feeding studies in rats, mice, dogs, and rabbits found no evidence of cancer induction upon histopathological examination of selected organs and tissues. The available studies tested mice, dogs, and rabbits with extreme intakes of uranium corresponding to radioactivity exposures of as high as 1.0×104 nCi/kg/day (3.7×105 Bq/kg/day) (1.5×104 mg U/kg/day) for 30 days (Maynard and Hodge 1949; Tannenbaum and Silverstone 1951) or rats and dogs at 8.2×103 nCi/kg/day (3×105 Bq/kg/day) (1.2×104 mg U/kg/day) for 2 years (Maynard and Hodge 1949; Maynard et al. 1953).
2.2.3 Dermal Exposure
2.2.3.1 Death
No deaths have been reported in humans as a result of dermal exposure to uranium.
Deaths have occurred in animals after dermal exposure to uranium compounds from both single and repeated exposures. Generally, the more water-soluble uranium compounds were the most toxic and the rabbit was the most sensitive species. Deaths were due to renal failure.
In a series of 4-hour exposures to uranium compounds followed by washing with detergent and a 30-day observation period, the lowest reported LD50 value was 28 mg U/kg as uranyl nitrate in an ethereal solution in New Zealand rabbits (Orcutt 1949). Calculated LD50 values for identical exposures to uranyl nitrate were 1,190 mg U/kg for guinea pigs and 4,286 mg U/kg for mice. Insufficient fatalities occurred to calculate an LD50 for rats, but the mortality curve fell between that of the rabbits and the guinea pigs. Deaths mainly occurred 5 to 7 days after exposure and were due to renal failure. Similar experiments with other uranium compounds in rabbits using a lanolin vehicle showed that water-soluble compounds (uranyl fluoride, uranium tetrachloride, uranium pentachloride) were the most toxic; the slightly soluble compounds (uranium trioxide, sodium diuranate, ammonium diuranate) had intermediate toxicity; and the
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water insoluble compounds (uranium tetrafluoride, uranium dioxide, uranium peroxide, triuranium octoxide) caused no deaths (Orcutt 1949).
Chemically induced renal failure caused 100% mortality in male Wistar rats after 5 daily exposures to 237 or 1,928 mg U/kg/day as uranyl nitrate hexahydrate or ammonium uranyl tricarbonate, respectively, applied in a water-Vaseline¨ emulsion (De Rey et al. 1983). A 60% mortality rate was also reported for other male Wistar rats that received daily applications of 1,965 mg U/kg as uranyl acetate dihydrate for 1–11 days. No deaths were reported for other Wistar rats similarly treated with 2,103 mg U/kg/day as ammonium diuranate or to an unspecified dose of uranium dioxide (De Rey et al. 1983).
Intermediate-duration dermal exposure in guinea pigs indicated that smaller repeated doses were better tolerated than a large single dose when the total exposure was the same. In a 4-week experiment where exposure was to 379 mg U/kg as uranyl nitrate for 3 days per week, 14% mortality was observed (Orcutt 1949). If the same cumulative dose (4,741 mg U/kg) had been given in a single application, 86% mortality would have been expected.
The LD50 values for each species and other LOAEL values for mortality from exposure to uranium through the dermal route are presented in Table 2-4.
2.2.3.2 Systemic Effects
No studies were located regarding systemic effects in humans following dermal exposure to uranium compounds for acute, intermediate, or chronic durations.
No studies were located regarding the respiratory, cardiovascular, gastrointestinal, hematological, musculoskeletal, hepatic, or endocrine effects of uranium in animals following acute-, intermediate-, or chronic-duration exposure; regarding the renal effects following intermediate- or chronic-duration exposure; regarding the dermal or body weight effects following chronic-duration exposure; or regarding ocular effects following acute- or chronic-duration exposure. The existing animal data on renal, dermal, and body weight effects are limited to acute- and intermediate-duration exposures.
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The highest NOAEL values and all reliable LOAEL values in each species and duration category for adverse systemic effects from chemical exposure to uranium by the dermal route are presented in Table 2-4.
Renal Effects. Rabbits, guinea pigs, rats, and mice dermally exposed to uranyl nitrate hexahydrate for 1 day showed proteinuria for up to 10 days, followed by recovery to control values. The degree of proteinuria did not correlate well with the applied dose of uranium. Rabbits had elevated blood NPN at doses over 270 mg U/kg. The animals that died from dermal exposure to uranium had microscopic renal damage typical of uranium poisoning. The kidneys of the animals that did not die were essentially normal, which may reflect repair of acute renal injury (Orcutt 1949). Chemically induced renal failure caused 100% mortality in male Wistar rats after 5 daily exposures to 237 or 1,928 mg U/kg/day as uranyl nitrate hexahydrate or ammonium uranyl tricarbonate, respectively, applied in a water-Vaseline¨ emulsion (De Rey et al. 1983). Deaths from renal failure were also reported in this study for male Wistar rats that received daily applications of 1,965 mg U/kg as uranyl acetate dihydrate for 1–11 days.
Dermal Effects. No human studies were located regarding the dermal effects of uranium; however, no dermal effects were reported in studies of uranium miners, millers, and processors.
In animal studies, application of 41 mg U/kg as uranium pentachloride to the shaved backs of New Zealand white rabbits resulted in mild skin irritation (Orcutt 1949). Dermally applied uranium was also damaging to the epidermis in other animal studies. Application of 56.4 mg U/kg as uranyl nitrate hexahydrate to another group of rabbits resulted in superficial coagulation necrosis and inflammation of the epidermis, while a dose of 4.2 mg U/kg as uranyl nitrate hexahydrate applied in single or multiple sites for 5 weeks resulted in severe dermal ulcers. No untreated controls were used in the 5-week study (Orcutt 1949). Moderate erythema was observed in male and female New Zealand white rabbits after single applications of 1.4 mg U/kg as uranyl nitrate hexahydrate to their uncovered clipped skins (Orcutt 1949). An applied dose of 2,670 mg U/kg as ammonium diuranate for 1–10 daily applications to the shaved backs of a group of rats resulted in mild lesions on the skin of the rats, while a dose of 237 mg U/kg as uranyl nitrate hexahydrate resulted in disrupted membranes in the cell, mitochondria, and cell nucleus, as revealed by transmission electron microscopy (TEM). Light microscopy revealed swollen and vacuolated epidermal cells and damage to hair follicles and sebaceous glands in the uranyl nitrate hexahydrate-treated animals (De Rey et al. 1983).
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No dermal effects were seen following application of a single dose of 618 mg U/kg as uranyl fluoride, 666 mg U/kg as uranium trioxide, 195 mg U/kg as sodium diuranate, 198 mg U/kg as ammonium diuranate, 410 mg U/kg as uranium peroxide, 458 mg U/kg as uranium dioxide, or 147 mg U/kg as triuranium octaoxide in 50% aqueous solution to the shaved skin of New Zealand white rabbits (Orcutt 1949). No dermal effects were observed on the shaved backs of New Zealand white rabbits to which a single dose of 98 mg U/kg as a 65% concentration of the uranium tetrafluoride in lanolin was applied (Orcutt 1949). Similarly, application of 3,929 mg U/kg as uranyl acetate dihydrate or 2,103 mg U/kg as ammonium uranyl tricarbonate in water-Vaseline¨ emulsion to a 3 cm2 shaved area of the uncovered backs of 20 male Wistar rats in 1–10 daily applications had no effect on the skin of the rats (De Rey et al. 1983).
Body Weight Effects. In animal studies, significant weight loss was reported in rats after the following dermal applications over a 3 cm2 area: 3,948 mg U/kg as uranyl nitrate hexahydrate, 3,929 mg U/kg as uranyl acetate dihydrate, 2,103 mg U/kg as ammonium uranyl tricarbonate, or 2,670 mg U/kg as ammonium uranate to rats for 1–10 days (De Rey et al. 1983). Weight loss was also observed after single applications of 660 or 689 mg U/kg as uranium tetrachloride to guinea pigs, 616 or 948 mg U/kg as uranyl nitrate hexahydrate to mice, 85 mg U/kg as uranyl nitrate hexahydrate to rats, and 43 mg U/kg as uranyl nitrate hexahydrate to rabbits (Orcutt 1949).
Uranium (4.2 mg U/kg/day) applied as uranyl nitrate hexahydrate to the clipped backs of New Zealand white rabbits for 5 weeks also induced significant weight loss that peaked at 10–15 days after beginning treatment (Orcutt 1949). However, in several other animal studies, no changes in body weight in New Zealand white rabbits were reported following single dermal applications of 618 or 804 mg U/kg as uranyl fluoride, 344 mg U/kg as uranium pentachloride, 666 mg U/kg as uranium trioxide or uranyl fluoride, 344 mg U/kg as uranyl pentachloride, 195 mg U/kg as sodium diuranate, 198 mg U/kg as ammonium diuranate, 410 mg U/kg as uranium peroxide, 458 mg U/kg as uranium dioxide, or 147 mg U/kg as triuranium octaoxide (Orcutt 1949).
2.2.3.3 Immunological and Lymphoreticular Effects
No information was located regarding the effects of uranium on the immunological and lymphoreticular system in humans and animals following dermal exposure for any duration.
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2.2.3.4 Neurological Effects
No studies were located for humans regarding neurological effects following dermal exposure to uranium compounds; however, such effects have not been observed in studies involving workers in uranium mining, milling, and production.
In animal studies, neurological signs observed in rabbits in a test in which single dermal doses of 1.4, 3, 6, 30, or 85 mg U/kg as uranyl nitrate hexahydrate were applied included irritability, hyperactivity, upset equilibrium, limb rigidity, and respiratory arrest at all doses tested (Orcutt 1949). The LOAEL value for this study is presented in Table 2-4.
2.2.3.5 Reproductive Effects
No studies were located for humans and animals that described reproductive effects following dermal exposure to uranium for any duration.
2.2.3.6 Developmental Effects
No studies were located regarding effects of uranium on development in humans or animals following dermal exposure for any duration.
2.2.3.7 Genotoxic Effects
No information was located regarding the toxicity of uranium to genetic material in humans or animals following dermal exposure for any duration of exposure. Other genotoxicity studies are discussed in Section 2.5.
2.2.3.8 Cancer
No information on the cancer effects in humans or animals following dermal exposure to uranium for all durations of exposure was located; however, such effects have not been observed in studies involving uranium mining, milling, and production.
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2.3 TOXICOKINETICS
Overview. Absorption of uranium is low by all exposure routes (inhalation, oral, and dermal). Absorption of inhaled uranium compounds takes place in the respiratory tract via transfer across cell membranes. The deposition of inhalable uranium dust particles in the lungs depends on the particle size, and its absorption depends on its solubility in biological fluids (ICRP 1994, 1996) Estimates of systemic absorption from inhaled uranium-containing dusts in occupational settings based on urinary excretion of uranium range from 0.76 to 5%. A comprehensive review of the available data for a pharmacokinetic model used lung absorption factors of 2% to 4% for 3 month old children and 0.2% to 2% for adults, based on compound absorbability (ICRP 1996). Gastrointestinal absorption of uranium can vary from
95%) that enters the body is not absorbed and is eliminated from the body via the feces. Excretion of absorbed uranium is mainly via the kidney. The case of Gulf War veterans who were exposed to depleted uranium from inhalation, ingestion, and wounds, showed average urinary excretion, 7 years post exposure, of 0.08 μg U/g creatinine, with the highest rates around 30 μg/g (McDiarmad et al. 1999b).
2.3.1 Absorption
2.3.1.1 Inhalation Exposure
The deposition of inhalable uranium dust particles in the various regions of the lungs (extrathoracic, tracheobronchial, and deep pulmonary or alveolar) depends on the size of the particles. Particles larger than 10 μm are likely to be transported out of the tracheobronchial region by mucocilliary action and
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swallowed. Particles that are sufficiently small to reach the alveolar region (#10 μm AMAD) may transfer rapidly or slowly into the blood, depending on the solubility of the uranium compound. According to the ICRP (1996), a more soluble compound (uranium hexafluoride, uranyl fluoride, uranium tetrachloride, uranyl nitrate hexahydrate) is likely to be absorbed into the blood from the alveoli within days and is designated inhalation Type F (fast dissolution). A less soluble compound (uranium tetrafluoride, uranium dioxide, uranium trioxide, triuranium octaoxide) is likely to remain in the lung tissue and associated lymph glands for weeks and is designated Type M (medium dissolution). A relatively insoluble compound (uranium dioxide, triuranium octaoxide) may remain in the lungs for years and is designated Type S (slow dissolution).
Analysis of excreta of active uranium mill crushermen exposed to ore dust indicated that 1–5% of uranium entering the lungs was absorbed systemically and excreted in the urine, and 95–99% was eliminated in the feces. Absorption could have taken place in the lungs or in the gastrointestinal tract from swallowed particles cleared from the lungs (Fisher et al. 1983). Uranium workers exposed to high levels of uranium dust had a very low lung burden of uranium, indicating that only a small fraction penetrates into the alveolar region (West and Scott 1969) and remains there without being cleared (or being very slowly cleared) via retrograde tracheobronchial mucus transport to the gastrointestinal tract, into lymph nodes, or dissolved into the circulating blood.
Estimates of absorption into the blood were derived from the excretion data of uranium mill workers (Wrenn et al. 1985). They estimated the daily mean absorption of inhaled uranium by mill workers at 24 μg U/day (0.34 μg U/kg for 70-kg reference man) based on measured excretion in feces and workplace ambient air concentrations. The absorption of uranium by these workers was estimated as 0.76% (range, 0.4–1.6%). Control subjects in a study of differential metabolism of 230Th, 234U, and 238U inhaled in uranium ore dust included 3 retired uranium mill workers (4–14 years since last employment as uranium ore crushermen), and 3 volunteers who lived in uranium milling communities but had no uranium work history. Two consecutive 24-hour urine and fecal collections were obtained and analyzed for 234U and 238U. The apparent total intakes of uranium of these individuals ranged from 11 to 18 μg U/day for the controls and from 5.3 to 71 μg U/day for the retirees. Although large compared to uranium intakes estimated for city dwellers, the uranium intakes of these individuals are not unreasonable because uranium in potable waters and locally grown foods tends to be higher in uranium mining and milling
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communities. The mean uranium absorption calculated for the controls (0.82%; range, 0.6–1%) was not significantly different from that calculated for the retired uranium workers (0.94%; range, 0.55–1.6%) (Wrenn et al. 1985).
Urinary excretion data was used to estimate the absorption of uranium by workers accidentally exposed to uranium hexafluoride (Fisher et al. 1990). Estimated airborne concentrations were 20 mg uranium hexafluoride/m3 for a 1-minute exposure and 120 mg uranium hexafluoride/m3 for a 60-minute exposure
(15.2 and 91 mg U/m3, respectively) (USNRC 1986). Initial intakes of workers involved in the accident ranged from 470 to 24,000 μg uranium.
Higher absorption of uranium occurred in animal studies using aerosols of purified uranium compounds. In these studies, as in human studies, the solubility of the uranium compound and the size of the inhaled particles determined absorption. Reported absorption of the inhaled dose was 18–40% in rats and 20–31% in guinea pigs for uranium hexafluoride (Leach et al. 1984) and 23% for uranium trioxide in dogs (Morrow et al. 1972).
2.3.1.2 Oral Exposure
Experimental studies in humans consistently show that absorption of uranium by the oral route is less than 5%. Reported fractional absorptions include a range of 0.005–0.05 (0.5–5%) in a group of four males ingesting 10.8 mg uranium in a soft drink (Hursh et al. 1969), less than 0.0025–0.04 in a group of 12 volunteers given drinking water high in uranium (Wrenn et al. 1989), and 0.005–0.05 in another drinking water study (Harduin et al. 1994). Similar results were obtained in dietary balance studies (Leggett and Harrison 1995; Spencer et al. 1990; Wrenn et al. 1989). A review of human data conducted by the ICRP determined that a fractional absorption of 0.02 for soluble compounds and 0.002 for insoluble compounds should be used in modeling the kinetics of dietary uranium in humans (ICRP 1995).
In animal studies, absorption generally increases with increasing solubility of the compound, being greatest for uranium ingested as uranyl nitrate hexahydrate, uranium hexafluoride or uranyl fluoride, about half as great for uranium tetroxide or uranium trioxide, and 1–2 orders of magnitude lower for uranium tetrachloride, triuranium octaoxide, and uranium tetrafluoride (ICRP 1995). Increased absorption of uranium has been demonstrated in neonatal rats and pigs (ICRP 1995). Fractional
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absorption in 2-day-old rats given uranyl nitrate was estimated as 0.01–0.07, two orders of magnitude greater than for adults (ICRP 1995).
Evidence from several animal studies showed that the amount of uranium absorbed from the gastrointestinal tract was about 1% (Harrison and Strather 1981; Larsen et al. 1984; LaTouche et al. 1987; Maynard et al. 1953; Sullivan 1980a). A range of gastrointestinal absorption rates of 0.038–0.078% has been estimated by others based on data from a 2-year study in which rats were fed diets containing 0.05–0.5% of soluble uranium compounds (uranyl fluoride or 0.5–2% of uranyl nitrate). The rate of absorption appeared to be independent of concentration of uranium in the diet (Wrenn et al. 1985). Absorption factors in rats that were exposed by gavage to doses of 233U-uranyl nitrate hexahydrate (where this anthropogenic radionuclide provided increased sensitivity without competition with natural isotopes) increased 3.4 times over normal in rats that were iron-deficient (Sullivan and Ruemmler 1988), doubled in rats that were fasted (Sullivan et al. 1986), and increased 3.6 times in neonates as compared to adults (Sullivan 1980b). Adult baboons (fed normally) absorbed about 0.5%, whereas fasted baboons absorbed an average of 4.5% (Bhattacharyya et al. 1989). Consistent with the results in baboons, fed and 24-hour fasted male B6CF1/ANL mice absorbed 0.069% and 0.80%, respectively (Bhattacharyya et al. 1989).
2.3.1.3 Dermal Exposure
Absorption of uranium through the skin has not been characterized in humans. Dermal absorption in animal models can be inferred from the appearance of toxicity in mice, rats, rabbits, and guinea pigs after dermal exposure to uranium compounds (Orcutt 1949). Absorption was also shown to occur through the conjunctival sac of the eye.
Electron microscopy and X-ray microanalytical methods showed that uranium as uranyl nitrate hexahydrate penetrated the stratum corneum within 15 minutes and accumulated in the intracellular space between the viable epidermis and the stratum corneum (De Rey et al. 1983). As is the case with inhalation and oral absorption, water solubility is an important determinant of absorption, and no penetration was observed with the insoluble compounds uranium dioxide, uranyl acetate, or ammonium diuranate. After 48 hours, uranium applied as uranyl nitrate was no longer found in the skin and toxicity developed, indicating that the uranium had been absorbed into the blood.
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2.3.2 Distribution
Absorbed uranium is found in all human tissues, but preferentially deposits in bone and kidney, regardless of the route of exposure (ICRP 1995, 1996). Although uranium also distributes significantly to liver, this organ is not a major repository for uranium; however, for modeling purposes, tissue contents are often normalized to liver concentration because the latter is reported in almost all studies of uranium biokinetics. The normal adult’s body burden is considered to be approximately 90 μg. It is estimated that about 66% of this total is in bone, 16% in the liver, 8% in the kidneys, and 10% in other tissues (ICRP 1979, 1995, 1996). It is not known if maternal bone stores of uranium (like those of calcium and lead) are mobilized during pregnancy and lactation. Uranium can cross the placenta after parenteral administration in animals; no information was located on distribution of uranium in breast milk for either humans or animals.
2.3.2.1 Inhalation Exposure
Autopsy data from individuals occupationally exposed to uranium indicates that bone is the primary site of long term retention of absorbed uranium (ICRP 1995). Inhalation exposure may also result in some retention of insoluble uranium particles in the lungs. An evaluation of the postmortem data from a uranium worker who had inhaled a total of 220 mg (147 pCi) uranium over a 3-year period found 11 μg (7 pCi) uranium in the lungs 13 years after the end of exposure. The total calculated dose equivalent from the inhaled uranium was 35 rem (0.35 Sv) (Keane and Polednak 1983).
In a comprehensive study of tissues from two long-time residents of New Mexico without known occupational exposure, the skeleton was the primary depot for uranium (Kathren 1997). Approximately 80 soft tissue samples and 90 bone samples were analyzed from each subject. The mean uranium concentrations in bone were 4.8 and 5.8 ng/g wet weight for the two subjects, respectively. Highest concentrations of uranium in soft tissues were in the tracheobronchial and other pulmonary related lymph nodes indicating uranium-bearing particulate clearance from the lungs. Concentrations in pulmonary lymph nodes ranged from 16–28 ng/g in one individual to 29–259 ng/g wet weight in the other.
Urinary excretion data were used in a kinetic model to estimate the maximum uranium kidney concentrations of workers accidentally exposed to uranium hexafluoride (Fisher et al. 1990). Initial intakes of workers involved in the accident ranged from 470 to 24,000 μg uranium. The model estimated the
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maximum kidney concentrations in the workers as ranging from 0.048 to 2.5 μg U/g in kidney tissue; renal toxicity was not observed in any of the workers (Fisher et al. 1990).
In animals, uranium that has been absorbed from the lungs leaves the blood very quickly for distribution to body tissues. The insoluble compounds (uranium tetrafluoride, uranium dioxide) were found to accumulate in the lungs and lymph nodes with the amount retained dependent on the exposure concentration and duration. In a continuous exposure study, more than 90% of the uranium retained at the end of the first year of exposure to a uranium dioxide aerosol was cleared by the end of the second year despite continued inhalation of uranyl nitrate. All of the uranium retained following one year of inhalation of uranyl hexafluoride was cleared by the end of the second year. For uranyl nitrate inhalation, no retention was found in the soft tissues. Uranium has also been shown to accumulate in the tracheobronchial lymph nodes, lungs, bones, and kidneys of rats, dogs, and monkeys exposed to uranium dioxide at 5 mg U/m3 for 1–5 years. Total radiation absorbed dose in dog lungs was around 600 rads (6 Gy). Up to 7,000 rads (70 Gy) were absorbed by monkey lymph nodes (Leach et al. 1973). In rats exposed to yellowcake, the U3O8 portion of the yellowcake cleared from the lung with a half-time of 110–240 days (Damon et al. 1984). Mice given inhaled doses of U3O8 equivalent to about 0.2 mg U/kg exhibited uranium tissue distribution (in μg/g tissue) as follows: lung, 6.05; liver, 0.051; spleen, 1.45; kidney, 0.536; tibia, 0.731; urine, 0.519; and feces, 2.20 (Walinder 1989). In an inhalation study using highly enriched uranium dioxide particles (92.8% 235U), rat lungs were found to clear the uranium particles at a rate of 0.28% per day over a period of 720 days. At 720 days postexposure, 82% of the uranium remained in the lungs and thoracic lymph nodes of the rats. The highest mass of extrapulmonary uranium dioxide was detected in rats sacrificed up to 11 days postexposure. This was mainly found in the intestinal tract and the carcass. The authors found that the pulmonary clearance rate of highly enriched uranium dioxide particles was about the same as the clearance rate for natural or unenriched uranium dioxide particles (Morris et al. 1990), as would be expected since they are the same chemical compound.
One site of deposition for the soluble compounds (uranyl nitrate, uranium tetrachloride, uranium hexafluoride) in animals was the skeleton, but accumulation was not seen in bone at levels below
0.25 mg U/m3 over a period of 2 years in rats exposed to soluble compounds (uranyl nitrate, uranium tetrachloride, uranium hexafluoride) in one study. The insoluble compounds (uranium hexafluoride, uranium dioxide) were found to accumulate in the lungs and lymph nodes after the inhalation exposure. For uranyl nitrate exposure, no retention was found in the soft tissues. Accumulation of uranium was also
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found in the skeleton (Stokinger 1953). The amount distributed in the skeleton has been reported to be 23–45% of the intake in dogs (Morrow et al. 1972); 28–78% in rats (Leach et al. 1984); and 34–43% in guinea pigs (Leach et al. 1984). A biological half-time of 150–200 days (Ballou et al. 1986) or 70 days (Morrow et al. 1982) in the skeleton has been reported following inhalation exposure to soluble uranium compounds (e.g., uranium hexafluoride).
A 5-year exposure of Beagle dogs and monkeys, and a 1-year exposure of rats, to 5.8 mg uranium dioxide/m3 (5.1 mg U/m3) as uranium dioxide dust (AMAD=1 μm) resulted in rapid lung buildup during the first few months, which approached maximal values of 2, 3.6, and 0.8 mg U/g in dogs, monkeys, and rats, respectively, at the end of year 1. Buildup in the tracheobronchial lymph nodes reached peak values in year 4 of 50–70 mg U/g in both dogs and monkeys. For each, the peak radiation dose rates reached 1.8 and 3.3 rads/week (0.018 and 0.033 Gy/week) to lungs, and 55 and 64 rads/week (0.55 and 0.65 Gy/week) to lymph nodes, while the total radiation dose for the 5 years approached 500 and 900 rads (5 and 9 Gy) to lungs and 10,000 rads (10 Gy) to the lymph nodes. Renal damage was not observed in either the dog or monkey, but fibrosis was found in the monkey lung and both necrosis and fibrosis were found in the dog and monkey lymph nodes. It was not clear whether the damage was chemically or radiologically induced, but the presence of lung and lymph node damage in the absence of renal effects was suggestive to the authors of long-term radiation damage (Leach et al. 1970). A reevaluation of the study data also showed a rapid accumulation of uranium in the lungs and tracheobronchial lymph nodes during the first few months of exposure. The accumulation in these organs was highest (0.8 mg/g in lungs and 1.5 mg/g in lymph nodes) at the end of 1 year of exposure. The uranium content in the lungs decreased with a half -time of approximately 480 days. In the lymph nodes, uranium depletion showed a trend similar to the lungs in dogs exposed for 2 and 5 years and a biphasic pattern in dogs exposed for 1 year. Comparatively low levels of uranium were found in the kidney, femur, liver, and spleen, and these decreased with time (Leach et al. 1973).
In other studies, no significant accumulation was found in the spleen or liver of rats, dogs, or guinea pigs (Ballou et al. 1986; Diamond et al. 1989; Leach et al. 1973, 1984; Morrow et al. 1972; Wrenn et al. 1987).
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2.3.2.2 Oral Exposure
Uranium levels have been measured in tissues from humans, with no occupational exposure where the source of uranium was assumed to be dietary and environmental.
In a comprehensive study of tissues from two long-time residents of New Mexico, the skeleton was the primary depot for uranium (Kathren 1997). Approximately 80 soft tissue samples and 90 bone samples were analyzed from each subject. The mean uranium concentrations in bone were 4.8 and 5.8 ng/g wet weight for the two subjects, respectively. Highest concentrations of uranium in soft tissues were in the tracheobronchial and other pulmonary related lymph nodes, indicating uranium-bearing particulate clearance from the lungs. Concentrations in pulmonary lymph nodes ranged from 16–28 ng/g in one individual to 29–259 ng/g wet weight in the other. An unexpectedly high concentration was found in the thyroid of one subject. In both subjects, uranium was widely distributed among the soft tissues; liver concentrations were lower than those in the kidney (approximately 0.1 ng/g and 0.9 ng/g wet weight, respectively).
The concentrations of uranium in human blood from New York City donors averaged 0.14 mg U/kg in both whole blood and red cells, compared to values ranging from <0.04 to 86 mg U/kg globally (Fisenne and Perry 1985). The median concentrations of uranium in the lungs, liver, kidneys, and vertebra from New York City residents among all age groups were reported to be 0.33, 0.13, 0.32, and 0.29 mg U/kg, respectively (Fisenne and Welford 1986). The concentration of uranium in human fat with no known occupational exposure was 0.6 ng/g (EPA 1985).
In an evaluation of two human skeletal tissues, it was observed that the sacrum contained the highest concentrations of 238U and 234U (4.9 mBq/g ash) (0.13 pCi/g ash) (0.20 μg/g ash). The concentration of 238U was lowest (0.1 mBq/g ash) (0.0027 pCi/g ash) (0.004 μg/g ash) in the right femur (Singh et al. 1987b). In the United Kingdom, the uranium concentration in wet bone was reported to be 3 mg U/kg (2 nCi U/kg) (Fisenne and Welford 1986).
Data on laboratory animals indicate that a substantial portion of uranium leaving the blood may initially distribute throughout soft tissues, but a few days after absorption or injection into the blood, most of the systemic content is found in the kidneys and skeleton (Bhattacharyya et al. 1989; ICRP 1995).
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In animals, a substantial fraction of plasma uranium is associated with the ultrafilterable low-molecularweight fraction, and the remainder is weakly associated with transferrin and other plasma proteins. Data on baboons indicate that 50% or more of the uranium in blood is associated with the red blood cells during the period 10–1,000 hours after injection. These data have been interpreted to mean that about 0.7% of the uranium leaving the plasma attaches to red blood cells and is returned to plasma with a halftime slightly greater than 1 day (ICRP 1995).
In animals, absorbed uranium is osteotropic, accumulating largely on the surface of all types of bone of the animals. Eventually, the uranium on the bone surface diffuses into the mineral portion of the bone. Autoradiography provides confirming evidence that, in the long term, uranium is a bone volume seeker (Wrenn et al. 1987). Kinetic models of uranium distribution predict that, for the short-term, uranium distributes to the bone surface and bone marrow while the deep bone is the long-term depot (Sontag 1986). These results suggest that the macro distribution of uranium in the human skeleton is not uniform.
In some ways, the skeletal behavior of uranium is quantitatively similar to that of alkaline earths. It is known that the uranyl ion (UO22+) exchanges with Ca2+ on the surfaces of bone mineral crystals, although it does not participate in crystal formation or enter existing crystals. The early distribution of uranium in different parts of the skeleton is similar to that of calcium. Uranium initially deposits on all bone surfaces but is most highly concentrated in areas of growth. Depending on the microscopic structure of the bone of each species, uranium on bone surfaces may gradually diffuse into bone volume; such diffusion has been observed in dogs but not in rats or mice. As with calcium, a substantial portion of uranium deposited in bone is lost to plasma by processes that occur more rapidly than bone resorption (see Section 2.3.5). In human subjects injected with uranium, an estimated 80–90% of the original skeletal deposition was lost from bone over the first 1.5 years (ICRP 1995).
In a study with female mice exposed orally in feed to uranyl nitrate hexahydrate at a dosage of 462 mg U/kg/day for 36–44 weeks, average uranium accumulation was 6 μg per pair of kidneys, 46 μg/g bone and 0–0.5 μg in whole liver, respectively. No significant organ accumulation was found for the lower dose levels (Tannenbaum and Silverstone 1951). Maximal concentrations of 77 μg per pair of kidneys and 216 μg/g in bone were estimated at 50 weeks in male mice that were orally exposed to uranyl nitrate hexahydrate at 925 mg U/kg/day for 48 weeks. One mouse with small kidneys showed levels of 395 μg/pair of kidneys and 1,440 μg/g bone (Tannenbaum and Silverstone 1951). Average uranium
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accumulation in the kidneys and bone of male mice exposed to uranyl fluoride orally at 452 mg U/kg/day for 28 weeks was 33 μg/pair of kidneys and 145 μg/g bone at 20 weeks (Tannenbaum and Silverstone 1951). Maximal concentrations of 6 μg/pair of kidneys at 50 weeks and 29 μg/g bone at 14 weeks were found in female mice given oral uranium tetrachloride at 978 mg U/kg/day for 48 weeks (Tannenbaum and Silverstone 1951).
The insoluble compounds of uranium accumulated to a lesser extent in tissues. Only small amounts of uranium were found in the kidneys (3–9 μg/pair of kidneys) of female mice that were exposed orally to uranium tetrafluoride at 4,437 mg U/kg/day for 48 weeks. No uranium was found in the bone (Tannenbaum and Silverstone 1951). Only small amounts of uranium were found in kidney (1–3 μg/pair of kidneys) of female mice that were exposed orally to triuranium octaoxide at 1,655 mg U/kg/day for 48 weeks. No uranium was found in the bone (Tannenbaum and Silverstone 1951).
2.3.2.3 Dermal Exposure
No studies were located regarding distribution of uranium after dermal exposure in humans or animals.
2.3.2.4 Other Routes of Exposure
Intravenously injected uranium is rapidly taken up by the tissues or excreted in the urine (ICRP 1995). Typically, 25% of intravenously injected uranium (as uranyl nitrate) remained in blood of human subjects after 5 minutes, 5% after 5 hours, 1% after 20 hours, and less than 0.5% after 100 hours although inter-subject variation was high (Bassett et al. 1948; Bernard and Struxness 1957). Measurements of systemic distribution of uranium made at autopsy in one terminally ill human given a single intravenous injection of uranium indicated that the skeleton, kidneys, and other soft tissues after 2.5 hours contained about 10, 14, and 6%, respectively, of the dose. Distribution data taken from another human subject 18 hours after a single intravenous injection uranium showed that the bones, kidneys, and other soft tissues contained about 4–13%, 6%, and 4%, respectively, of the amount injected. At 566 days post-injection, uranium distribution in the skeleton, kidneys, and other soft tissues declined to about 1.4, 0.3, and 0.3%, respectively.
The distribution of uranium metal implanted in muscle has been investigated in rats (Pellmar et al. 1999a). In these experiments, pellets of depleted uranium were implanted into the gastrocnemius muscle
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and uranium levels were measured in kidney, muscle, liver, spleen, brain, serum and bone at 1 day and at 6, 12, 18 months after implantation. Within 1 day uranium was measurable in kidney and bone but not in the other tissues. At later time points, significant amounts of uranium were found in the other tissues, although levels were always highest in the kidney and bone.
2.3.3 Metabolism
Uranium is usually found in compounds which can be metabolized and recomplexed to form other compounds. In body fluids, tetravalent uranium is likely to oxidize to the hexavalent form followed by formation of uranyl ion. Uranium generally complexes with citrate, bicarbonates, or protein in plasma (Cooper et al. 1982; Dounce and Flagg 1949; Stevens et al. 1980). The stability of the carbonate complex depends on the pH of the solution, which will differ in different parts of the body (BEIR IV 1988). The low-molecular-weight bicarbonate complex can be filtered at the renal glomerulus, and be excreted in urine at levels dependent on the pH of the urine. The uranium bound to the protein (primarily transferrin) is less easily filtered and is more likely to remain in blood. In the blood, the uranyl ion binds to circulating transferrin, and to proteins and phospholipids in the proximal tubule (Wedeen 1992).
2.3.4 Elimination and Excretion
Two-thirds of uranium, intravenously injected as uranyl nitrate in human subjects was typically excreted in urine in the first 24 hours. Approximately 10% more was excreted over a period of 5 days. Fecal excretion accounted for less than 1% of the excretion (ICRP 1995).
2.3.4.1 Inhalation Exposure
In a study of 7,231 uranium workers, the urinary concentration of uranium ranged from 5 μg/L in 4,556 workers to more than 100 μg/L in 32 workers. Samples were taken weekly over a 6-year period. Among a control group of 600 non-uranium workers, none had urinary uranium concentrations that exceeded 40 μg/L. The author concluded that urinary uranium concentrations greater than 100 μg/L are definitely indicative of recent absorption, and that pathological albuminuria is rare, except when the urinary uranium concentration exceeds 1,000 μg/L. Albuminuria, when seen, was transient, and did not persist (Butterworth 1955).
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Urinary excretion in crushermen (about 0.2 nCi/day [7 Bq/day][0.3 mg/day]) is about 1/100th of fecal excretion (about 13.5 nCi/day [500 Bq/day][20 mg/day]). The activity of 234U in urine was slightly higher than that of 238U. Active crushermen excreted higher levels of 234U, 238U, and 230Th than retired crushermen or controls (Fisher et al. 1983). Most of the inhalation doses of female employees at the Oak Ridge plant were excreted in the feces, indicating that ciliary action in the lungs, followed by fecal excretion, was an important mechanism of body clearance (West and Scott 1969).
Zhao and Zhao (1990) reported on the excretion of uranium in an occupationally exposed worker. A 23-year old man who weighed 60 kg, dressed in protective clothing, mask, and gloves, was accidentally exposed to pure uranium tetrafluoride powder for 5 minutes. The uranium tetrafluoride powder cloud was reported to contain natural uranium. Urinary excretion was reported as 112 μg/L or
156.8 μg in the first 24 hours, gradually increasing through post-accident day 60 and returning to normal at about post-accident day 1,065. The total urinary excretion of uranium through day 1,065 was calculated to be 86.7 mg. The excretion data was used to calculate total absorption and kidney content by use of a kinetic model (ICRP 1979). The kidney content on post-accident day 1 was reported as 804.2 μg or approximately 2.6 μg/g of kidney.
The biological half-time of uranium dioxide in human lungs (occupational exposure) at German fuel fabrication facilities was estimated to be 109 days. Body burden measurements of uranium taken from 12 people who handled uranium oxides for 5–15 years were used for this determination. Twice a year for 6 years, a urinalysis was conducted on workers exposed to uranium. In vivo lung counting was performed on the last day before and the first day after a holiday period. Levels of uranium in feces were measured during the first 3 days and the last 3 days of a holiday period and the first 3 days after the restart of work. For some employees, the levels of uranium in feces was measured during 3–4 days one-half year after the holiday period (Schieferdecker et al. 1985).
In animals, most of absorbed uranium is excreted in urine. Inhaled larger particles ($10 μm) are transported out of the respiratory system by mucocilliary action, then swallowed, and eliminated in the feces (Ballou et al. 1986; Downs et al. 1967; Morrow et al. 1982). Deposition sites of inhaled aerosols, and hence the clearance kinetics, are determined in part by particle size of the inhaled particles. As the AMAD increases, the amount deposited in the upper respiratory tract increases, and the amount deposited in the deep respiratory tracts of the lungs decreases. This study used both 232U and 233U dusts. The 233U dust deposition in the upper respiratory tract increased from 21 to 62% of the total
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amount of dust deposited with increasing particle size; deposition in the deep lung decreased from 22 to 7% with increasing particle size. The 232U dust deposition in the upper respiratory tract increased from 10 to 32% with increasing particle size; deposition in the deep lung decreased from 23 to 9% with increasing particle size. The differences were less marked for 232U dust, presumably, because the particle size was much more uniform than that for the 233U dust. A large amount of the initial lung burden was preferentially cleared via the feces following clearance from the upper respiratory tract to the gastrointestinal tract (higher fecal excretion with higher AMAD) by mucocilliary action. Urinary excretion was 25–50% of initial lung burden on day 1; less with larger particles. By day 7, 25–80% of the uranium uptake was cleared in urine; most of the uranium was eliminated in the feces (Ballou 1986). In one study with rats, most of the inhaled uranium, as uranium dioxide, was excreted in the urine. In dogs, less than 10% was excreted in feces (presumably cleared by mucocilliary action) (Cooper et al. 1982). About 60% of the retained uranium, as uranyl nitrate hexahydrate (Ballou et al. 1986), uranium hexafluoride (Leach et al. 1984), and uranium trioxide (Morrow et al. 1982), was excreted in urine within 1 day in other studies with rats, dogs, and guinea pigs. Most of the retained uranium in rats exposed via intratracheal intubation with uranium dioxide or uranyl nitrate hexahydrate was excreted in the urine. Less than 10% was excreted in feces (presumably cleared by mucocilliary action) (Cooper et al. 1982). The fraction of insoluble compounds (uranium tetrafluoride, uranium dioxide) retained in the lungs and lymph nodes was independent of the exposure concentration. More than 90% of the uranium retained at the end of the first year of exposure to uranium dioxide was cleared by the end of the second year despite continued exposure to uranyl nitrate hexahydrate. All of the uranium retained following 1 year of exposure to uranium tetrafluoride was cleared by the end of the second year. For uranyl nitrate hexahydrate exposure, no retention was found in the soft tissues (Stokinger 1953).
Once deposited in the lungs, uranium compounds clear from the various biological compartments by solubility. The ICRP lung model recognizes three clearance classification types: F, M, and S. Type F compounds (uranium hexafluoride, uranyl fluoride, uranium trioxide, possibly uranium tetrafluoride, possibly triuranium octaoxide) show 100% absorption with a half-time of 10 minutes. Type M compounds (uranyl nitrate, ammonium diuranate, possibly uranium tetrafluoride, possibly triuranium octaoxide) show 10% absorption with a half-time of 10 minutes, and 90% with a half-time of 140 days, and about 70% of the material in the alveoli eventually reached body fluid. Type S compounds (uranium dioxide) show 0.1% absorption with a half-time of 1 minute, and 99.9% with a half-time of 7,000 days,
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and 10% of that deposited in the alveoli reaches body fluid (ICRP 1996). The half-time of uranium in the lungs has also been calculated to be 1–5 days for soluble compounds like uranyl nitrate hexahydrate in rats (Ballou et al. 1986), ammonium diuranate in hamsters (Stradling et al. 1984), and uranyl fluoride in dogs (Morrow et al. 1982). It is longer for the less soluble uranium dioxide: 141–289 days in rats (Downs et al. 1967) and 480 days in dogs (Leach et al. 1973). In the kidney, uranium selectively accumulates in the proximal tubule with a biological half-time of about 1 week (Wedeen 1992). The halftime of uranyl fluoride in the kidneys has been reported to be 2–5 days in rats (Diamond et al. 1989) and 9 days in dogs. In dogs, less than 1% of the uranium remained in the kidneys after 30 days (Morrow et al. 1982).
2.3.4.2 Oral Exposure
The available evidence on the excretion of ingested uranium suggests that most ($95%) is excreted in the feces, and the remainder in urine (Wrenn et al. 1985). Urinary uranium excretion rates from nonoccupationally exposed persons in 3 villages near uranium mining and refining facilities and a control village in Japan ranged from <0.02–0.24 mg U/day per person and <0.02–0.04 mg U/day per person, respectively (Masuda 1971). The half-time in the kidneys has been estimated to be 1–6 days for 99% of the uranium in the kidneys and 1,500 days for the remainder (ICRP 1979). Most of the uranium doses, given as 900 mL of water containing 90 pCi (3.3 Bq) 234U and 90 pCi (3.3 Bq) 238U (180 pCi or 6.6 Bq uranium) to drink over a period of 6 hours, was excreted in feces within 2 days (Singh and Wrenn 1987). Four volunteers who ingested 10.8 mg of uranium mixed with Coca Cola excreted the uranium in both feces and urine over a 25-day period (Hursh et al. 1969). Urinary excretion after oral exposure is generally low and has been estimated as 2% of total excretion (Spencer et al. 1990).
Animal studies have shown that most ingested uranium (99%) is not absorbed in rats, but is eliminated in the feces without being cycled through the bile. In rats, most of the absorbed uranium leaves the body within a few days in urine; half is excreted in 2–6 days (Durbin and Wrenn 1975), and 98% within 7 days (Sullivan 1986). About 95% of the uranium in the kidneys of rats is excreted in urine within 1 week, and very little remains in any other organ (LaTouche et al. 1987; Sullivan 1980a, 1986).
Data from parenteral studies provide further indication that uranium retention in animal kidneys is described by a 2-compartment exponential curve. Reported biological half-times for the compartments
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are 2 and 50–60 days (Diamond et al. 1989), 2 and 13 days (Bentley et al. 1985), or 3 and 103 days (Wrenn et al. 1986).
2.3.4.3 Dermal Exposure
No studies were located describing the excretion of uranium following dermal exposure in humans or animals.
2.3.5 Physiologically Based Pharmacokinetic (PBPK)/Pharmacodynamic (PD) Models
Physiologically based pharmacokinetic (PBPK) models use mathematical descriptions of the uptake and disposition of chemical substances to quantitatively describe the relationships among critical biological processes (Krishnan et al. 1994). PBPK models are also called biologically based tissue dosimetry models. PBPK models are increasingly used in risk assessments, primarily to predict the concentration of potentially toxic moieties of a chemical that will be delivered to any given target tissue following various combinations of route, dose level, and test species (Clewell and Andersen 1985). Physiologically based pharmacodynamic (PBPD) models use mathematical descriptions of the dose-response function to quantitatively describe the relationship between target tissue dose and toxic endpoints.
PBPK/PD models refine our understanding of complex quantitative dose behaviors by helping to delineate and characterize the relationships between: (1) the external/exposure concentration and target tissue dose of the toxic moiety and (2) the target tissue dose and observed responses (Andersen and Krishnan 1994; Andersen et al. 1987). These models are biologically and mechanistically based and can be used to extrapolate the pharmacokinetic behavior of chemical substances from high to low dose, from route to route, between species, and between subpopulations within a species. The biological basis of PBPK models results in more meaningful extrapolations than those generated with the more conventional use of uncertainty factors.
The PBPK model for a chemical substance is developed in four interconnected steps: (1) model representation, (2) model parameterization, (3) model simulation, and (4) model validation (Krishnan and Andersen 1994). In the early 1990s, validated PBPK models were developed for a number of toxicologically important chemical substances, both volatile and nonvolatile (Krishnan and Andersen
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1994; Leung 1993). PBPK models for a particular chemical substance require estimates of the chemical substance-specific physicochemical parameters, and species-specific physiological and biological parameters. The numerical estimates of these model parameters are incorporated within a set of differential and algebraic equations that describe the pharmacokinetic processes. Solving these differential and algebraic equations provides the predictions of tissue dose. Computers then provide process simulations based on these solutions.
The structure and mathematical expressions used in PBPK models significantly simplify the true complexities of biological systems. This simplification, however, is desirable if the uptake and disposition of the chemical substance(s) is adequately described because data are often unavailable for many biological processes and using a simplified scheme reduces the magnitude of cumulative uncertainty. The adequacy of the model is, therefore, of great importance and, thus, model validation is important.
PBPK models improve the pharmacokinetic extrapolation aspects of the risk assessment process, which seeks to identify the maximal (i.e., safe) levels for human exposure to chemical substances (Andersen and Krishnan 1994). PBPK models provide a scientifically based means to predict the target tissue dose of chemicals in humans who are exposed to environmental levels (for example, levels that might occur at hazardous waste sites) based on the results of studies where doses were higher or were administered in different species. Figure 2-4 shows a conceptualized representation of a PBPK model. The overall results and individual PBPK models are discussed in this section in terms of their use in risk assessment; tissue dosimetry; and dose, route, and species extrapolations.
The ICRP (1994, 1996) developed a Human Respiratory Tract Model for Radiological Protection which contains respiratory tract deposition and clearance compartmental models for inhalation exposure that may be applied to uranium. The ICRP (1995) also developed a biokinetic model for human oral exposure that applies to uranium. Two other compartmental models (Fisher et al. 1991; Sontag et al. 1986) are also described below. The National Council on Radiation Protection and Measurement (NCRP) has also developed a respiratory tract model for inhaled radionuclides (NCRP 1997). At this time, the NCRP recommends the use of the ICRP model for calculating exposures for radiation workers and the general public. Readers interested in this topic are referred to NCRP Report No. 125; Deposition, Retention and Dosimetry of Inhaled Radioactive Substances (NCRP 1997). In the appendix to the report, NCRP provides the animal testing clearance data and equations fitting the data which supported the development of the human model.
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Human Respiratory Tract Model for Radiological Protection (ICRP 1994).
Deposition. The ICRP has developed a deposition model for behavior of aerosols and vapors in the respiratory tract. It was developed to estimate the fractions of radioactivity in breathing air that are deposited in each anatomical region. ICRP provides inhalation dose coefficients which can be used to estimate the committed equivalent and effective doses to organs and tissues throughout the body based on a unit intake of radioactive material. The model applies to three levels of particle solubility, a wide range of particle sizes (approximately 0.0005–100 μm in diameter), and parameter values can be adjusted for various segments of the population (e.g., sex, age, level of physical exertion). This model also allows one to evaluate the bounds of uncertainty in deposition estimates. Uncertainties arise from natural biological variability among individuals and the need to interpret some experimental evidence that remains inconclusive. It is applicable to particles containing uranium, but was developed for a wide variety of radionuclides and their chemical forms.
The ICRP deposition model estimates the amount of inhaled material that initially enters each compartment (see Figure 2-5). The model was developed with 5 compartments: (1) the anterior nasal passages (ET1); (2) all other extrathoracic airways (ET2) (posterior nasal passages, the naso- and oropharynx, and the larynx); (3) the bronchi (BB); (4) the bronchioles (bb); and (5) the alveolar interstitium (AI). Particles deposited in each of the regions may be removed from each region and redistributed either upward into the respiratory tree or to the lymphatic system and blood by different particle removal mechanisms.
For extrathoracic deposition, the model uses experimental data, where deposition is related to particle size and airflow parameters, and scales deposition for women and children from adult male data. Similarly to the extrathoracic region, experimental data served as the basis for lung (bronchi, bronchioles, and alveoli) aerosol transport and deposition. A theoretical model of gas transport and particle deposition was used to interpret data and to predict deposition for compartments and subpopulations other than adult males. Table 2-5 provides reference respiratory values for the general Caucasian population under several levels of activity.
Respiratory Tract Clearance. This portion of the model identifies the principal clearance pathways within the respiratory tract. The model was developed to predict the retention of various radioactive materials. Figure 2-6 presents the compartmental model and is linked to the deposition model
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(Figure 2-5) and to reference values presented in Table 2-6. Table 2-6 provides clearance rates and deposition fractions for each compartment for insoluble particles. The table provides rates of insoluble particle transport for each of the compartments, expressed as a fraction per day and also as half-time. ICRP also developed modifying factors for some of the parameters, such as age, smoking and disease status. Parameters of the clearance model are based on human evidence for the most part, although particle retention in airway walls is based on experimental data from animal experiments.
The clearance of particles from the respiratory tract is a dynamic process. The rate of clearance generally changes with time from each region and by each route. Following deposition of large numbers of particles (acute exposure), transport rates change as particles are cleared from the various regions. Physical and chemical properties of deposited material determine the rate of dissolution and as particles dissolve, absorption rates tend to change over time. By creating a model with compartments of different clearance rates within each region (e.g., BB1, BB2, BBseq), the ICRP model overcomes problems associated with time-dependent functions. Each compartment clears to other compartments by constant rates for each pathway.
Particle transport from all regions is toward both the lymph nodes and the pharynx, and a majority of deposited particles end up being swallowed. In the front part of the nasal passages (ET1), nose blowing, sneezing, and wiping remove most of the deposited particles. Particles remain here for about a day. For particles with AMADs a few micrometers or greater, the ET1 compartment is probably the largest deposition site. A majority of particles deposited at the back of the nasal passages and in the larynx (ET2) are removed quickly by the fluids that cover the airways. In this region particle clearance is completed within 15 minutes.
Ciliary action removes deposited particles from both the bronchi and bronchioles. Though it is generally thought that mucocilliary action rapidly transports most particles deposited here toward the pharynx, a fraction of these particles are cleared more slowly. Evidence for this is found in human studies. For humans, retention of particles deposited in the lungs (BB and bb) is apparently biphasic. The “slow” action of the cilia may remove as many as half of the bronchi- and bronchiole-deposited particles. In human bronchi and bronchiole regions, mucus moves more slowly the closer to the alveoli it is. For the faster compartment it has been estimated that it takes about 2 days for particles to travel from the bronchioles to the bronchi and 10 days from the bronchi to the pharynx. The second (slower) compartment is assumed to have approximately equal fractions deposited between BB2 and bb2 and both with clearance
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half-times estimated at 20 days. Particle size is a primary determinant of the fraction deposited in this slow thoracic compartment. A small fraction of particles deposited in the BB and bb regions is retained in the airway wall for even longer periods (BBseq and bbseq).
If particles reach and become deposited in the alveoli, they tend to stay imbedded in the fluid on the alveolar surface or move into the lymph nodes. The one mechanism by which particles are physically resuspended and removed from the AI region is coughing. For modeling purposes, the AI region is divided into 3 subcompartments to represent different clearance rates, all of which are slow.
In the alveolar-interstitial region, human lung clearance has been measured. The ICRP model uses 2 halftimes to represent clearance: about 30% of the particles have a 30-day half-time, and the remaining 70% are given a half-time of several hundred days. Over time, AI particle transport falls and some compounds have been found in lungs 10–50 years after exposure.
Absorption into Blood. The ICRP model assumes that absorption into blood occurs at equivalent rates in all parts of the respiratory tract, except in the anterior nasal passages (ET1), where no absorption occurs. It is essentially a 2-stage process, as shown in Figure 2-7. First, there is a dissociation (dissolution) of particles; then the dissolved molecules or ions diffuse across capillary walls and are taken up by the blood. Immediately following dissolution, rapid absorption is observed. For some elements, rapid absorption does not occur because of binding to respiratory-tract components. In the absence of specific data for specific compounds, the model uses the following default absorption rate values for those specific compounds that are classified as Types F (fast), M (medium), and S (slow):
C For Type F, there is rapid 100% absorption within 10 minutes of the material deposited in the BB, bb, and AI regions, and 50% of material deposited in ET2. Thus, for nose breathing, there is rapid absorption of approximately 25% of the deposit in ET and 50% for mouth breathing. Type F uranium compounds include uranium hexafluoride, its mixture with uranyl fluoride, uranyl nitrate (which can behave as Type M), pure uranium trioxide, and uranium tetrafluoride (which can behave at Type M).
C For Type M, about 70% of the deposit in AI reaches the blood eventually. There is rapid absorption of about 10% of the deposit in BB and bb, and 5% of material deposited in ET2. Thus, there is rapid absorption of approximately 2.5% of the deposit in ET for nose breathing, and 5% for mouth breathing. Type M compounds include unpure uranium trioxide, uranyl nitrate (which can behave as Type F), ammonium diuranate, uranium oxtaoxide (which can behave as Type S), and uranium tetrafluoride (which can behave as Type F).
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For Type S, 0.1% is absorbed within 10 minutes and 99.9% is absorbed within 7,000 days, so there
is little absorption from ET, BB, or bb, and about 10% of the deposit in AI reaches the blood
eventually. Type S compounds include uranium dioxide and uranium octaoxide (which can behave
as Type M).
Biokinetic Model for Uranium (ICRP 1995). The ICRP biokinetic model for uranium is based on the generic model structure for alkaline earth elements described in Publication 67 (ICRP 1993, as cited in ICRP 1995). Uranium (as the UO22+ ion) is similar to calcium (Ca2+) with regard to skeletal kinetics. Some transfer rates in the biokinetic model for uranium are equated with bone formation rates. The early behavior of uranium in human circulation is represented reasonably well by treating plasma as a uniformly mixed pool, where uranium is removed at a rate of 35 d-1 (ICRP 1995) and where a soft tissue compartment (ST0) is in relatively rapid exchange with plasma (see Figure 2-8). Compartment ST0 is assumed to receive 30% of uranium leaving plasma and to have a removal half-time of 2 hours (from ST0 to plasma). ICRP assumed that 1% of uranium leaving the circulation (or 0.7% leaving plasma) deposits in red blood cells (ICRP 1995). The removal half-time from red blood cells to plasma is assumed to be 2 days.
Urinary excretion of uranium is assumed to arise from uranium moving directly from plasma to the urinary bladder contents. Approximately 60% of uranium leaves the blood directly to the bladder and another 12% is retained temporarily in the renal tubules before excretion. The liver is assumed to consist of two compartments, Liver 1 and Liver 2. The liver receives an estimated 1.5% of uranium leaving the blood, with over 90% returning to circulation.
Little direct information on the kinetics of uranium in children exists. Age-specific deposition of uranium in the skeleton is assumed to be proportional to the deposition of the alkaline earth elements. The rate of removal from deep bone is assumed to be the same as the age-specific bone turnover rate. Because children have higher amounts of uranium taken up by bone, deposition in soft tissues and excreta are likely lower in children than for adults.
Sontag (1986) Pharmacokinetic Model. An extended multicompartmental model (see Figure 2-9) describing the kinetic behavior of uranium (absorption, distribution, and excretion as a function of time) in the organs of male and female rats was developed using data taken from experiments performed on 13-month-old male and female Sprague-Dawley rats intravenously injected with 1.54 mCi/kg (57 kBq/kg) 233U-uranyl citrate and sacrificed at 7, 28, 84, 168, or 336 days after injection.
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The model is composed of 10 compartments. These 10 compartments are connected by 17 linear transfer coefficients using 21 parameters. The whole system describes the flux of compounds between a central compartment (the blood) and outer compartments which connect with the central compartment only. The 10 compartments are labeled blood, bone 1, bone 2, liver 1, liver 2, kidney 1, kidney 2, residual 1, residual 2, and excretion. The organs are divided into two compartments; one compartment represents the short term and one represents the long term. For example, the short-term compartment for the bone is the bone surface and bone marrow, and the long-term compartment is the deep bone. In the liver, the short-term compartment is assumed to be the lysosomes, and the long-term compartment is assumed to be the telolysosomes. Separation of these organs into two components helps to account for the reabsorption and rapid excretion. Using the symbols BP=blood, EC=excretion, B1=bone 1, L1=liver 1, K1=kidney 1, R1=residual 1, B2=bone 2, L2=liver 2, K2=kidney, and R2=residual 2, the calculated transfer coefficients for this model are shown in Table 2-7.
Parallel evaluations produced 2 different values (ranges) for each of the 21 parameters. The maximum fractions of uranium in various compartments were as follows: bone, 0.0710 or 0.0735; liver, 0.0160 or 0.0146; kidney, 0.1777 or 0.4789; residual compartment, 0.0358 or 0.0481; and excretion compartment, 0.6995 or 0.3849 (if no back transfer to the blood compartment occurred). The time at which the maximum amount of the uranium in the organ is reduced to one-half is 0.0009 or 0.0013 days in the blood, 165 or 93 days in the bone, 6 or 7 days in the liver, 11 or 5 days in the kidney, and 5 or 6 days in the residual compartment. The cumulative radiation absorbed dose in the organ 365 days after injection of 56.6 kBq/kg body weight was 0.0002 or 0.0004 Gy to blood, 0.730 or 1.29 Gy to bone, 0.0268 or 0.0308 Gy to liver, 1.32 or 1.77 Gy to kidney, and 0.0061 or 0.0076 Gy to residual compartment. The ratio of single injection/continuous intake calculated for the same dose 1 year after the first injection was
0.018 or 0.003 to blood, 0.619 or 0.812 to bone, 0.422 or 0.355 to liver, 0.256 or 0.231 to kidney,
0.726 or 0.585 to residual compartment, and 1.024 or 1.023 to excretion compartment (Sontag 1986).
Fisher et al. (1991) Biokinetic Model A modified biokinetic model for uranium was developed for inhaled soluble uranium based on human data from an accidental release of uranium hexafluoride in Oklahoma. Urinary excretion data from 31 exposed workers were used to test two previously published compartmental models for inhalation exposure to uranium (ICRP 1979; Wrenn et al. 1989). Urinary uranium was measured periodically for 2 years following the accident. Statistical analysis showed that the Wrenn et al. (1989) model produced a better fit to the excretion data than the ICRP (1979) model.
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Parameters of the (Wrenn et al. 1989) model were then modified to further improve the fit to the workers excretion data. Changing the retention half-time in the kidney from 15 days to 6 days and the clearance half-time in the lung from 0.5 days to 0.03 days optimized the fit of the model to the experimental data. The model may be summarized with the following 5-term exponential equation:
y(t) ' 1.5e &2.77t % 0.028e &0.116t % 0.0069e &0.0347t % (4.8 x 10&7e &0.000462t) % 3.2 x 10&6e &0.000139t
u
where, yu(t) is fractional daily uranium excretion rate at t days after intake; the excretion constants in the 5 exponents corresponding to compartments with retention half-times of 0.25, 6, 20, 1,500, and 5,000 days.
The model was used to estimate uranium intakes; uranium burdens in the lungs, kidneys, and bones; and effective dose equivalent for each worker in the accident. Initial intakes of workers involved in the accident ranged from 470–24,000 μg uranium. The model estimated the maximum kidney concentrations in the workers as ranging from 0.048 to 2.5 μg U/g kidney tissue, renal toxicity was not observed in any of the workers (Fisher et al. 1990, 1991).
Based on this same data base, the NRC determined that the maximum uranium dose equivalent of workers on-site was 28 mrem (0.28 mSv). The maximum uranium dose equivalent of off-site individuals was
1 mrem (0.014 mSv). However, these radiological doses were small compared to the background radiation level of 106 mrem/year (1.06 mSv/year) in the area from which the data were collected (USNRC 1986).
2 MECHANISMS OF ACTION

2.4.1 Pharmacokinetic Mechanisms
On the average, a given amount of an ingested uranium compound appears to be less toxic than the same amount of an inhaled uranium compound (Maynard and Hodge 1949; Stokinger et al. 1953). This finding may be partly attributable to the relatively low gastrointestinal absorption of uranium compounds. Only 0.1–6% of even the more soluble uranium compounds are absorbed in the gastrointestinal tract (Harrison
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and Strather 1981; Hursh et al. 1969; ICRP 1979; Larsen et al. 1984; LaTouche et al. 1987; Leggett and Harrison 1995; Maynard et al. 1953; Sullivan 1980a; Wrenn et al. 1985, 1988b). The ICRP (1995) recommends a gastrointestinal absorption reference fraction of 0.02 for uranium ingested in relatively soluble form and 0.002 for insoluble compounds. On the basis of the toxicity of different uranium salts in animals, it was concluded that the relatively more soluble salts (uranyl nitrate hexahydrate, uranyl fluoride, uranium tetrachloride, uranium pentachloride) were most toxic, the slightly soluble compounds (uranium trioxide, sodium diuranate, ammonium diuranate) were of intermediate toxicity, and the insoluble compounds (uranium tetrafluoride, uranium dioxide, uranium tetrachloride, triuranium octaoxide) were nontoxic (Orcutt 1949).
In inhalation exposures, uranium compounds are usually inhalable aerosols. Thus, particle size plays a vital role in tissue dose. Particles larger than 5 μm AMAD are likely to be transported out of the tracheobronchial region by mucocilliary action and swallowed into the gastrointestinal tract, where absorption is minimal (ICRP 1979). The less soluble compounds (uranium trioxide, uranium tetrafluoride), designated Type M by the ICRP (1995), are more likely to remain in the lung tissue and associated lymph glands for weeks. The relatively insoluble compounds (uranium dioxide, triuranium octaoxide), designated Type S by the (ICRP 1995), are likely to remain in the lungs for years (Eidson 1994). This retention of uranium in the lung can lead to a significant pulmonary radiation dose.
In addition, the sequestration patterns of the different uranium compounds are important determinants for the target organ chemical and radiological toxicities of these compounds. The site of deposition for the soluble uranium compounds (uranyl nitrate, uranium tetrachloride, uranium hexafluoride) is the bone, while the insoluble compounds (uranium hexafluoride, uranium dioxide) accumulate in the lungs and lymph nodes (Stokinger 1953).
2.4.2 Mechanisms of Toxicity
The dual modes of uranium chemical and radiological toxicity are not usually separately identifiable by end point. The renal and respiratory effects from exposure of humans and animals to uranium are usually attributed to the chemical properties of uranium, while the theoretically potential excess cancers are usually attributed to the radiation properties of this substance. Although the net effects on the lungs and kidneys have been suggested to be a cooperative action of the chemical and radiation properties, with a
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complementary mechanism of action, this relationship has not been demonstrated experimentally (Ballou et al. 1986; Dockery et al. 1993; Dungworth 1989; Filippova et al. 1978; Leach et al. 1984; Spiegl 1949; Spoor and Hursh 1973; Stokinger et al. 1953). UNSCEAR has considered that limits for natural (and depleted) uranium in drinking water (the most important source of human exposure) should be based on the chemical toxicity rather than on a radiological toxicity, which has not been observed in either humans or animals (UNSCEAR 1993; Wrenn et al. 1985).
The most sensitive indicator of uranium toxicity to mammals, and perhaps humans, is nephrotoxicity. While acute high level exposure to uranium compounds can clearly cause nephrotoxicity in humans (Pavlaikis et al. 1996; Zhao and Zhao 1990), the evidence for similar toxicity as the result of long-term lower level occupational exposures is equivocal. Epidemiologic studies have not noted an increase in deaths from urogenital or renal diseases (Brown and Bloom 1987; Checkoway et al. 1988; Dupree et al. 1987; Lundin et al. 1969; Polednak and Frome 1981), and follow-up studies have failed to identify significant damage to human kidneys following occupational exposure to uranium (Eisenbud and Quigley 1955; Hursh and Spoor 1973; Luessenhop et al. 1958), for which regulatory limits are set to prevent damage. A recent comparison of autopsy kidney tissue samples revealed no differences between 7 uranium workers and 6 referents with no known exposure to uranium (Russell et al. 1996). One epidemiologic study provided evidence of nephrotoxicity following occupational exposure to uranium. Nephrotoxicity, indicated by β2-microglobulinuria and aminoaciduria due to decreased tubular reabsorption, was reported in a group of 39 male uranium mill workers exposed for more than a year to uranium concentrations exceeding the occupational standard of 3.7 Bq/m3 (currently 5 Bq/m3 [0.2 mg/m3]) by #8-fold. Cement workers were used as controls in this study (Thun et al. 1985).
Many animal studies have shown that inhalation, oral, or dermal exposure to uranium results in kidney damage. The damage was histologically manifested as glomerular and tubular wall degeneration. Ultrastructural analysis showed damage to the endothelial cells in the glomerulus, such as loss of cell processes, and reduction in the density of the endothelial fenestrae (Avasthi et al. 1980; Haley 1982; Haley et al. 1982; Kobayashi et al. 1984). In the terminal segments of the proximal convoluted tubules, there was a loss of the brush border, cellular vacuolization, and necrosis. Tubular reabsorption of solutes was disrupted. Functionally, this process led to a disruption of the tubular solute reabsorption and to a decrease in the filtration rate of the glomerulus, as assessed by creatinine or inulin clearance or by proteinuria (Bentley et al. 1985; Blantz 1975; Leach et al. 1973; Morrow et al. 1982). Excessive urinary excretion of protein,
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glucose, enzymes, or amino acids such as catalase or alkaline phosphatase are additional indicators of uranium-induced renal pathology (Maynard et al. 1953) by inhalation exposure (Bentley et al. 1985; Diamond et al. 1989; Haley et al. 1982; Leach et al. 1984; Maynard et al. 1953; Morrow et al. 1982).
A mechanism involving bicarbonate activity in the kidneys has been proposed for uranium-induced renal toxicity. Uranium is usually combined with either bicarbonate or a plasma protein in the blood. In the kidneys, uranium is released from bicarbonate and is free to combine to form complexes with phosphate ligands and proteins in the tubular wall to cause damage. Uranium is not tightly bound and is released again within a few days. Within a week following exposure, uranium is largely cleared from the kidneys, and the tubules begin to regenerate. Although the regenerated epithelium has histological differences from its normal state, it is often difficult to detect histological signs of kidney damage a month after exposure because all remaining functional damage is subtle. An alternative mechanism through which uranium exerts its renal toxicity has been suggested by the results of a study conducted with rabbit kidney cells in vitro. In this study, uranyl nitrate hexahydrate inhibited both sodium transport-dependent and independent ATP utilization and mitochondrial oxidative phosphorylation in the renal proximal tubule. Ouabain-insensitive adenosine triphosphatase (ATPase) activity exhibited the greatest sensitivity to uranyl nitrate hexahydrate and was significantly inhibited at submillimolar concentrations (Brady et al. 1989). Perhaps both of these activities combine to cause renal damage. In addition, because uranium is a predominantly alpha-emitting radionuclide, current theories on cellular necrosis by high-LET alpha radiation imply a contributory role to the cellular degenerative nephrotoxic changes (BEIR 1980, 1988, 1990; Filippova et al. 1978; Sanders 1986; UNSCEAR 1982, 1986, 1988).
Most studies of respiratory diseases reported for uranium involve noncancerous alveolar epithelium damage in type II cells. These changes are characterized by interstitial inflammation of the alveolar epithelium leading eventually to emphysema or pulmonary fibrosis in acute exposures or to hyperplasia, hypertrophy, and transdifferentiation (metaplasia) in chronic exposures (Cooper et al. 1982; Dungworth 1989; Stokinger 1981; Wedeen 1992). However, the lack of significant pulmonary injury in most inhalation animal studies indicates that other potentially toxic contaminants such as inhalable dust particles, radium, or radon may contribute to these effects.
Large doses of ionizing radiation have the actual or theoretical potential of being carcinogenic, teratogenic, and mutagenic. Since uranium has a low specific activity but emits high LET alpha particles
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that are densely ionizing along their track length, studies have been conducted to determine if uranium can produce these effects in humans and animals. The 4 to 8 MeV alpha particles from uranium travel through 40–70 μm in soft tissue, incrementally transferring their kinetic energy to the series of atoms and molecules with which they interact along their short, straight paths. Consequently, only structures within this range from the site of the deposition of uranium may be affected. If a DNA molecule is intersected and damaged without resulting in cell death, a range of theoretical effects can result. DNA has been found to be the most radiosensitive biological molecule, and ionizing radiation has been observed to damage individual chromosomes. The main result from low level ionizing radiation exposure is DNA damage or fragmentation. Viable cells repair the damage, but repair errors can result which produce gene mutations or chromosomal aberrations. Such events may result in such highly rare events as carcinogenesis or teratogenesis, but there is currently no evidence for radiation mutagenesis in humans. Chromosomal aberrations following large radiation doses have been demonstrated in humans and in research animals, showing that ionizing radiation can both initiate and promote carcinogenesis, and interfere with reproduction and development. Cancer is a well-known effect of ionizing radiation exposure, but it has never been associated with exposure to uranium. Likewise, no genetic changes due to radiation have ever been observed in any human population exposed at any dose (BEIR 1980, 1988, 1990; Leach et al. 1970; Morris et al. 1990; Muller et al. 1967; Otake and Schull 1984; Sanders 1986; Stokinger et al. 1953; UNSCEAR 1982, 1986, 1988). For these reasons, UNSCEAR has stated that limits for natural (and depleted) uranium in drinking water (the most important source of human exposure) should be based on the chemical toxicity rather than on a hypothetical radiological toxicity in skeletal tissues, which has not been observed in either humans or animals (Wrenn et al. 1985). The EPA also used chemical toxicity as the basis for their 20 μg/L interim drinking water limit for uranium published in 1991 (currently withdrawn).
2.4.3 Animal-to-Human Extrapolations
Kidney damage and respiratory disease are the most significant health effects in animals from the metallotoxicity of uranium. Because the biological systems through which these effects are mediated are common to both animals and humans (Brady et al. 1989; Cooper et al. 1982; Dungworth 1989; Stokinger 1981; Wedeen 1992), it is reasonable that animals are appropriate surrogates for humans in this regard. This assumption is consistent with evidence in humans for respiratory (Kathren and Moore 1986, Waxweiler et al. 1981a) and renal (Bernard and Struxness 1957; Fisher et al. 1991; Kathren and Moore 1986; Luessenhop et al. 1958; Thun et al. 1985; USNRC 1986; Waxweiler et al. 1981a; Zhao and Zhao
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1990) effects. The data from these studies support the assumption of biological similarity in the renal toxicity of uranium in animals and humans. Nevertheless, a considerable uncertainty is associated with animal-to-human extrapolation regarding the renal toxicity of uranium exposure because the renal toxicity of animals varies with species.
2.5 RELEVANCE TO PUBLIC HEALTH
Overview.
Uranium is an alpha-emitting, radioactive, heavy metal that occurs naturally in the earth's crust at an average concentration of about 2 ppm (approximately 1 pCi/g). Uranium exists in several isotopic forms. The most toxicologically important forms are anthropogenic 232U and 233U and naturally occurring 234U, 235U, and 238U. Uranium isotopes decay by alpha emission. 238U decays through 16 radioactive progeny, including 234U, to reach stable lead-206 (206Pb), while 235U decays through 13 radioactive progeny to reach stable 207Pb. This profile discusses the chemical and radiological health effects of isotopes of uranium (natural, enriched, and depleted) and the various compounds in which uranium is usually found. The health effects of daughter isotopes (radium and radon) are addressed in other toxicological profiles (consult the ATSDR toxicological profiles for radium and radon for more information regarding these radionuclides).
Naturally occurring uranium is an isotopic mixture containing a large percentage of 238U and very small percentages of 234U and 235U, by mass. The industrial process called enrichment is used to increase the percentage of 235U and decrease the percentage of 238U in natural uranium. This results in a continuum of additional isotope mixtures in which the percentage of 235U is either larger (enriched uranium) or smaller (depleted uranium) than that of natural uranium. Natural uranium consists of 99.284% 238U, 0.711% 235U, and 0.005% 234U by weight and has a very low specific activity (0.68 μCi/g). Uranium enrichment for commercial nuclear energy produces uranium that contains about 3% 235U; this is called 3% enriched uranium. Uranium enrichment for other purposes, including nuclear weapons production, can produce uranium containing as much as 97.3% 235U and having a higher specific activity (.50 μCi/g). Depleted uranium is the byproduct of the enrichment process. Depleted uranium has even less specific activity
(0.33 μCi/g) than natural uranium.
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Uranium is present in the body at very low or trace concentrations and is not known to be an essential element. Human intakes are constant through very small amounts of natural uranium in food and water, and even smaller amounts in air. The following anthropogenic activities increase the potential for human exposure to uranium: mining, milling, and handling uranium; processing uranium ore end products (uranium dioxide, uranium hexafluoride); producing nuclear energy and nuclear weapons; producing phosphate fertilizers from phosphate rocks that contain much higher-than-average levels of uranium; and improperly disposing of wastes. Occupational exposure to airborne uranium ore dust occurs in uranium mines and mills and in processing plants. Typically, uranium represents only 0.2–5% by weight of the ore.
The deposition of inhaled dust particles in the lungs depends on the particle size and the absorption depends on the solubility of the compound. Very small particles, on the order of 1 μm AMAD, are deposited in the alveolar region or deep lung spaces. As particle size increases above 2–3 μm AMAD, there is an increasing likelihood of deposition in the tracheobronchial region. Dust particles that have deposited are rapidly transported out of the tracheobronchial region by mucociliary action and swallowed. The more soluble compounds are more likely to be absorbed into the blood at the alveolar level within days. Ingested uranium that has been cleared from the lungs by mucocilliary action and swallowed is only partly absorbed into the blood. This is true even for the more common soluble salts (uranium hexafluoride, uranyl fluoride, uranium tetrachloride, uranyl nitrate hexahydrate). Uranium is usually found in compounds that can break down and recomplex to form other compounds. In body fluids, tetravalent uranium is likely to oxidize to the hexavalent form, followed by formation of the uranyl ion. Uranium generally complexes with citrate, bicarbonates, or protein in plasma.
According to the ICRP (1995), the more soluble compounds (uranium hexafluoride, uranyl fluoride, uranium tetrachloride, uranyl nitrate hexahydrate) are more likely to be absorbed into the blood from the alveoli within days and are assigned to inhalation Type F (fast dissolution). The less soluble compounds (uranium tetrafluoride, uranium dioxide, uranium trioxide, triuranium octaoxide) are more likely to remain in the lung tissue and associated lymph glands for weeks and are designated Type M (medium dissolution). The relatively insoluble compounds (uranium dioxide, triuranium octaoxide) may remain in the lungs for years and are designated Type S (slow dissolution). The ICRP (1995) recommends the following absorption factors for humans for inhaled compounds that subsequently enter the gastrointestinal tract: 2% for soluble compounds and 0.2% for less soluble compounds.
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The main site of long-term retention for soluble uranium compounds (uranyl nitrate, uranium tetrachloride, uranium dioxide) is the bone, while the inhaled insoluble compounds (uranium tetrafluoride, uranium dioxide) that are deposited in the deep respiratory tract tend to accumulate in the lungs and pulmonary lymph nodes.
Ingested uranium is excreted mostly in the feces; urinary excretion is generally low. The biological halftimes of soluble uranium compounds (uranium hexafluoride, uranyl fluoride, uranium tetrachloride, uranyl nitrate hexahydrate) are estimated in days or weeks; those of the less soluble compounds (uranium tetrafluoride, uranium dioxide, triuranium octaoxide) are estimated in years. No information is currently available on the excretion of dermally absorbed uranium. Transdermally absorbed uranium is expected to behave identically to uranium compounds absorbed through the lungs and the gastrointestinal tract.
Because the specific activities of natural and depleted uranium are low, no remarkable noncancerous radiological health hazard is expected (and none has been observed) from exposure to natural and depleted uranium. The results of the available studies in humans and animals are consistent with this conclusion. According to the BEIR IV report, if uranium’s radiation were carcinogenic in humans, the most likely carcinogenic effect in humans would be bone sarcoma. However, even highly-enriched uranium has not been found to produce cancer, including that of the bone, in exposed humans. Evidence from animal studies suggests adverse effects reported from such exposures include damage to the epithelium of the lungs (fibrosis) and cardiovascular abnormalities (friable vessels).
The chemical action of all isotopes and isotopic mixtures of uranium are identical, regardless of the specific activity, because chemical action depends only on chemical properties. Thus, the chemical toxicities of natural, depleted, and enriched uranium are identical. Current evidence from animals studies suggests that the toxicity of uranium is mainly due to its chemical damage to kidney tubular cells, leading to nephritis.
Evidence also suggests that the toxicity of uranium varies according to the route of exposure and to its compounds. This finding may be partly attributable to the relatively low gastrointestinal absorption of uranium compounds. Only 1,000 fold) above any plausible human exposure either in the workplace or at hazardous waste sites (Leach et al. 1984; Spiegel 1949; Maynard and Hodge 1949; Orcutt 1949). On the basis of the available data, exposure to environmental uranium or to uranium at levels found at hazardous waste sites will not be lethal to humans.
Systemic Effects.
Respiratory Effects. General damage to pulmonary structures, usually noncancerous alveolar epithelium damage of type II cells, can occur upon inhalation of insoluble reactive chemicals such as some uranium compounds (uranium tetrafluoride, uranium dioxide, uranium trioxide, triuranium octaoxide). In acute
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exposures, pulmonary damage may be limited to interstitial inflammation of the alveolar epithelium leading eventually to emphysema or pulmonary fibrosis (Cooper et al. 1982; Dungworth 1989; Saccomanno et al. 1982; Stokinger 1981; Wedeen 1992). In studies of the pulmonary effects of airborne uranium dust in uranium miners (Dungworth 1989; Waxweiler et al. 1983) and in animals (Filippova et al. 1978; Leach et al. 1984; Spiegl 1949; Stokinger et al. 1953), the respiratory diseases reported were aggravated by the insoluble aerosol particles (mine dust) to which these miners were exposed because most of the noncancerous respiratory diseases reported in these studies were consistent with toxicity of inhalable dust particles other than uranium, such as silica (Dockery et al. 1993). This hypothesis is supported by the lack of respiratory diseases in laboratory animal models exposed to aerosols of uranium compounds in the absence of other aerosols (Maynard and Hodge 1949; Maynard et al. 1953; Stokinger et al. 1953; Gilman et al. 1998a; Gilman et al. 1998b; Gilman et al. 1998c). Reports of workers in the uranium-processing industry do not show increased deaths due to diseases of the respiratory system related to exposure to uranium (Brown and Bloom 1987; Cragle et al. 1988; Polednak and Frome 1981; Scott et al. 1972). Respiratory effects reported in workers acutely exposed to uranium hexafluoride were caused by hydrogen fluoride, a potent lung irritant and a spontaneous by-product of uranium hexafluoride (Kathren and Moore 1986; USNRC 1986).
Studies in humans that provide evidence for the radiotoxicity of uranium to the lungs were equivocal and unreliable for use in assessing uranium-specific hazards. The subjects in these studies were also concurrently exposed to known pulmonary toxicants and radiological agents (e.g., radon progeny), as well as to silica dust, which was identified as the etiological agent for silicosis (Dupree et al. 1987; Hadjimichael et al. 1983). Inhalation studies with animals regarding the association of respiratory disease to uranium exposure per se were equivocal (Cross et al. 1981a, 1981b, 1982; Leach et al. 1970, 1973, 1984; Morrow et al. 1982; Stokinger et al. 1985). No studies were found that reported respiratory effects in animals following oral or dermal exposure to uranium compounds. Thus, no adverse pulmonary effects from human exposure to uranium per se at or near hazardous waste sites are likely. However, prolonged exposure to high levels of insoluble uranium dust, as may occur with uranium miners, millers, and processors, or accidental exposure to high levels of soluble uranium aerosols, especially uranyl fluoride, may damage the lungs by chemical action. These conditions are unlikely at hazardous waste sites; therefore, respiratory effects are unlikely.
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Cardiovascular Effects. No reliable studies were identified that associated exposure to uranium with cardiovascular effects in humans. The available studies in animals (rats, mice, guinea pigs, and rabbits) found no adverse cardiovascular effects in animals following inhalation or oral exposures to uranium (Dygert 1949c; Gilman et al. 1998a, 1998b, 1998c; Maynard and Hodge 1949; Stokinger et al. 1953). Although a study in rats that used single intratracheal instillation of 90% enriched soluble uranium salts reported dystrophied blood vessels and enlarged hearts (Filippova et al. 1978), human exposures to such high specific-activity radionuclides at hazardous waste sites are unlikely. No studies were located that reported cardiovascular effects in animals following dermal exposure to uranium compounds. Therefore, no cardiovascular effects are likely from human exposure to environmental levels or to levels expected at or near hazardous waste sites.
Gastrointestinal Effects. A case report of a 5-minute accidental occupational exposure of a male worker in China to fumes of uranium tetrafluoride described signs and symptoms of gastrointestinal distress. These signs and symptoms included loss of appetite, abdominal pain, diarrhea, tenesmus, and pus and blood in the stool (Zhao and Zhao 1990). No gastrointestinal effects were seen in animals given unenriched uranium nitrate in doses as high as 664 mg U/kg/day for 2 years (Gilman et al. 1998a, 1998b, 1998c; Maynard et al. 1949). Gastrointestinal effects are not likely following exposure to uranium at hazardous waste sites.
Hematological Effects. The available human studies (Archer et al. 1973b; Eisenbud and Quigley 1955; Vich and Kriklava 1970) provide no clear evidence that uranium exposure can cause hematological effects in humans. Although the available animal studies provide evidence that very high exposure to uranium compounds may cause disruptions in the blood (Cross et al. 1981b; Dygert 1949b, 1949c, 1949d; Leach et al. 1970, 1973; Ortega et al. 1989a; Pozzani 1949; Roberts 1949; Rothermel 1949; Rothstein 1949b, 1949c, 1949d; Spiegl 1949; Stokinger et al. 1953), others provide evidence of no detectable hematological disturbances (Gilman et al. 1998a, 1998b, 1998c). Adverse effects on the blood are not expected health outcomes from exposure to uranium at the levels found at hazardous waste sites.
Musculoskeletal Effects. No studies have reported effects of uranium on the musculoskeletal system in humans following inhalation, oral, or dermal exposure for any duration. Laboratory animal studies support a lack of toxicological effects on the musculoskeletal system after oral exposures (Gilman et al. 1998a, 1998b, 1998c).
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Hepatic Effects. No reliable studies were located regarding the effects of uranium on the liver of humans following inhalation, oral, or dermal exposure for any duration. One case report does exist which documents hepatotoxicity in one male who drank 15 g of uranyl acetate (Pavlakis et al. 1996). Uranyl acetate is water soluble and would likely be more quickly absorbed from the gastrointestinal tract than the more insoluble forms of uranium. The patient suffered from increasing liver dysfunction, as evidenced by increased serum ALT, AST, and GGK. Since no liver biopsy sample was obtained, it is difficult to elaborate further on other liver changes that may have occurred. This individual had a history of drug abuse, which may have predisposed him to hepatic toxicity. The liver injury appeared temporary, with no residual signs of hepatotoxicity 6 months after ingestion.
No studies were located regarding the effects of uranium on the liver of animals following dermal exposure for any duration. No indications of liver damage were reported in several animal studies (Dygert 1949c; Pozzani 1949; Rothstein 1949c; Stokinger et al. 1953). However, inhalation exposure to relatively high concentrations of uranium compounds has resulted in mild liver disturbances, although the etiology is not clear. These disturbances include increased bromosulfalein retention, indicative of impaired biliary function, in a chronic-duration inhalation study in dogs (Stokinger et al. 1953); increased urinary catalase, moderate fatty livers, and a slight decrease in hepatic lactate content in rabbits (Roberts 1949; Rothstein 1949c, 1949d); and focal hepatic necrosis in rats (Dygert 1949a; Roberts 1949). Oral exposure of animals to uranium compounds was also accompanied by indications of mild liver damage (Gilman et al. 1998a, 1998c). These indications included microhemorrhagic foci in the liver of rats (Domingo et al. 1987); liver congestion, minimal hepatic lesions, mild degeneration, or fatty infiltration in dogs (Maynard and Hodge 1949); and increased lysosomal activity in dogs (Ortega et al. 1989a). No changes were seen in other dog studies in which the animals were given doses as high as 7,859 mg U/kg/day as the relatively insoluble uranium acetate dihydrate for 30 days (Maynard and Hodge 1949). Anisokaryosis, vesiculation, increased portal density, and perivenous vaculolation were observed in rats (Gilman et al. 1998a), while accentuation of zonation, variation in hepatocellular nuclear size, nuclear pyknosis, and excessive cytoplasmic vacuolization were observed in rabbits (Gilman et al. 1998c). The data presented here suggested that uranium is a hepatotoxicant by the inhalation and oral exposure routes in both humans (limited data set) and laboratory animals. Uranium disrupts general hepatocellular function and cellular permeability; however, no mechanism for these effects has been identified from any of these studies. On the basis of the available data, effects on the liver can occur as a result of human exposure to uranium compounds. However, human and animal studies indicate that the
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liver is an order of magnitude less sensitive than the kidney by either inhalation (Dygert 1949a, 1949d; Pozanni 1949; Rothstein 1949c; Stokinger 1953) or oral (Gilman et al. 1998a, 1998c; Maynard and Hodge 1949; Pavlikis et al. 1996) routes, for all exposure durations. Thus, it is highly unlikely that exposure to uranium compounds near hazardous waste sites could result in liver damage.
Renal Effects. Uranium has been identified as a nephrotoxic metal, exerting its toxic effect by chemical action mostly in the proximal tubules in humans and animals. There is sufficient information with high exposures to both soluble and insoluble uranium to permit the conclusion that uranium has a low order of metallotoxicity in humans in view of the high levels to which the subjects were exposed. The negative findings regarding renal injury among workers exposed to insoluble compounds are particularly significant in view of the high levels of exposure reported (Eisenbud and Quigley 1955).
Several epidemiologic studies found no increased deaths in uranium workers due to renal disease (Archer et al. 1973a, 1973b; Brown and Bloom 1987; Checkoway et al. 1988; Polednak and Frome 1981). Also, case studies showed that workers accidentally exposed to high levels of uranium did not have renal damage even up to 38 years postexposure (Eisenbud and Quigley 1956; Kathren and Moore 1986). However, one study on the kidney function of uranium mill workers chronically exposed to soluble uranium revealed renal tubular dysfunction as manifested by mild proteinuria, aminoaciduria, and a dose-related clearance of β2-microglobulin relative to that of creatinine. Serum β2-microglobulin was also elevated in the serum of 22 of the 23 workers tested. The incidence and severity of these nephrotoxic signs correlated with the length of time that the uranium workers had spent in the yellowcake (insoluble) drying and packaging area (Saccomanno et al. 1982; Thun et al. 1985). The data from this study were indicative of reduced protein resorption in the proximal renal tubules consistent with the observed renal toxicity of uranium in animals. Two case reports of accidental occupational exposures to high concentrations of both soluble and insoluble uranium by inhalation or dermal routes described clinical findings of a decreased glomerular filtration rate as manifested by decreased urinary output and significantly elevated urinary proteins, nonprotein nitrogen, amino acid nitrogen/creatinine, and phenolsulfonpthalein. Renal function rapidly returned to normal in days (Zhao and Zhao 1990). Acute nephrotoxicity was attributed to a large oral intake of uranyl acetate (Pavlakis et al. 1996). Similarly, the results of animal studies indicated that nephrotoxicity is the most consistent and sensitive adverse effect following inhalation (Dygert 1949a, 1949b, 1949c, 1949d; Filippova et al. 1978; Gilman et al. 1998a, 1998b, 1998c Leach et al. 1970, 1973, 1984; Pozzani 1949; Roberts 1949; Rothermel 1949; Rothstein
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1949a, 1949c, 1949d; Spiegl 1949; Sprague 1949; Stokinger et al. 1953), oral (Domingo et al. 1987; Gilman et al. 1998a, 1998b, 1998c; MacDonald-Taylor 1992; Maynard and Hodge 1949; Ortega et al. 1989a), or dermal (De Rey et al. 1983; Orcutt 1949) exposures to uranium compounds. These nephrotoxic effects are consistent with the metallotoxic action of uranium in the kidneys (Goodman 1985). The pathogenesis of the kidney damage in animals indicates that regeneration of the tubular epithelium occurred following discontinuation of exposure to uranium (Bentley et al. 1985; Dygert 1949b; Maynard and Hodge 1949; Pozzani 1949; Rothermel 1949; Rothstein 1949c; Spiegl 1949; Stokinger et al. 1953). Thus, exposure to the soluble compounds of uranium at or near hazardous waste sites could result in kidney damage. Measurement of uranium in air, soil, and water at or near the site is necessary to predict the likelihood of renal effects.
The following MRLs have been calculated for exposure to uranium based on kidney effects:
an intermediate-duration MRL of 8×10-3 mg/m3 for inhalation exposure to insoluble compounds of uranium is based on renal tubule lesions in dogs (Rothstein 1949b);
an intermediate-duration MRL of 4×10-4 mg/m3 for inhalation exposure to soluble compounds of uranium is based on renal tubule lesions in dogs (Rothstein 1949a);
a chronic-duration MRL of 3×10-4 mg/m3 for inhalation exposure to soluble compounds of uranium is based on renal tubule lesions in dogs (Stokinger et al. 1953);
an intermediate-duration MRL of 2×10-3 mg/kg/day for oral exposure to soluble compounds of uranium is based on renal tubule lesions in rabbits (Gilman et al. 1998b);

See the MRL discussion earlier in this section and in the MRL Worksheets in Appendix A for further details on the derivation of these MRLs.
Endocrine Effects. No endocrine effects were reported in humans following inhalation, oral, or dermal exposure to uranium compounds. No endocrine effects were reported in most of the available studies of animals following inhalation or oral exposure to uranium compounds (Gilman et al. 1998a, 1998b, 1998c; Ortega et al. 1989a; Maynard and Hodge 1949; Stokinger et al. 1953). A later study by Gilman et al. (1998a) using uranyl nitrate also identified endocrine organ changes that were limited to multifocal reductions in follicular size, increased epithelial cell height, and decreased amounts and densities of colloid in the thyroids of male but not female rats. Endocrine effects are not a significant health concern to individuals living at or near hazardous waste sites containing uranium.
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Dermal Effects. No significant dermal effects were observed in animals given inhalation or oral doses of uranium compounds (Spiegl 1949; Stokinger et al. 1953). However, dermal application of uranium compounds resulted in mild skin irritation, severe dermal ulcers, or superficial coagulation necrosis and inflammation of the epidermis in rabbits (Orcutt 1949). Dermal application resulted in swollen, vacuolated epidermal cells and damage to hair follicles and sebaceous glands in rats (De Rey et al. 1983). The effects or symptoms of acute dermal exposure to ionizing radiation included erythema (redness of the skin) and epilation (loss of hair) (Upton 1993). The alpha particle emitted by uranium will not penetrate the dead keratinized outer layer of the skin, so there is minimal concern for dermal effects from skin contact with uranium. Dermal effects were not seen in studies of uranium miners, millers, and processors. The observed skin damage reported in animals dermally exposed to excessive quantities of uranium compounds is not expected to occur in human exposures at hazardous waste sites. Such exposures, if they occur, are expected to be at or less than the levels at which uranium miners, millers, and processors are exposed (levels at which no attributable dermal health effects were reported).
Ocular Effects. No ocular effects attributable to uranium exposure were reported in the available human studies. In animal studies, dogs exposed to 13 mg U/m3 as uranium hexafluoride for 30 days exhibited encrusted eyes and conjunctivitis prior to death. However, these signs were considered nonspecific indications of poor health by the investigators of the study (Spiegl 1949). Consequently, no ocular effects are expected from human exposure to uranium compounds.
Body Weight Effects. Body weight loss was not reported in any of the human studies regarding inhalation, oral, or dermal exposure to uranium compounds. Similarly, the available studies in animals that evaluated this end point did not find any significant changes following inhalation exposure to uranium compounds (Cross et al. 1981b; Dygert 1949c, 1949d; Leach et al. 1970, 1973; Pozzani 1949; Rothermel 1949; Rothstein 1949a, 1949b, 1949c, 1949d; Spiegl 1949; Stokinger et al. 1953). Similar lack of body weight effects were found in rats in 28- and 91-day using uranium nitrate in the drinking water (Gilman et al. 1998a). The initial or reversible loss of body weight observed in animals exposed to high concentrations of uranium in the diet in acute-, intermediate-, and chronic-duration studies was accompanied by decreased food consumption due to taste aversion (Maynard and Hodge 1949, 1953; Tannenbaum and Silverstone 1951). This initial effect reversed, and the animals returned to their normal body weight as normal food intake resumed. Thus, the changes in body weight seen in such studies may be due more to reduction in food consumption due to bad taste than to uranium-specific toxicity. This
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effect may not be relevant to humans. It is more likely that significant weight loss in rabbits following application of excessive dermal doses (as high as 1,917 mg U/kg) (Orcutt 1949) may be a response to exceeding the maximally tolerated dose in these animals and consequently overwhelming the physiological mechanisms in these species (Orcutt 1949). Therefore, no significant effect on body weight is expected from human exposure to uranium compounds at or near hazardous waste sites.
Metabolic Effects. No studies were located regarding the metabolic effects in humans or animals following acute-, intermediate-, or chronic-duration inhalation, oral, or dermal exposure to uranium and uranium compounds. Consequently, it is not known whether human exposure to uranium and uranium compounds could result in adverse metabolic effects; however, such effects would not be anticipated, based on the absence of endocrine effects.
Other Systemic Effects. No studies were located that reported other systemic effects in humans or animals following inhalation, oral, or dermal exposure to uranium. Consequently, no other systemic effects are expected from human exposure to uranium compounds at or near hazardous waste sites.
Immunological and Lymphoreticular Effects. No adverse immunological or lymphoreticular effects were reported in human studies following exposure to uranium through the inhalation, oral, or dermal route for any duration (Brown and Bloom 1987; Checkoway et al. 1988; Cragle et al. 1988; Keane and Polednak 1983; Polednak and Frome 1981; Vich and Kriklava 1970). Similarly, no significant uranium-induced immunological or lymphoreticular changes were observed in animals exposed to uranium for acute, intermediate, or chronic durations (Filippova et al. 1978; Gilman et al. 1998a, 1998b, 1998c; Leach et al. 1970, 1973; Malenchenko et al. 1978; Maynard et al. 1953; Stokinger et al. 1953; Tannenbaum and Silverstone 1951). Sinus hyperplasia of the spleen was noted in rats in one 91-day uranium nitrate drinking water study (Gilman et al. 1998a). No significant immunological or lymphoreticular injury is expected from human exposure to uranium compounds at or near hazardous waste sites.
Neurological Effects. Although no neurological functions were evaluated in most of the available human studies, no damage to structures of the central or peripheral nervous system and no overt neuropathology were reported in humans following exposure to natural or enriched uranium compounds by the inhalation, oral, or dermal route (Brown and Bloom 1987; Carpenter et al. 1988; Cragle et al. 1988;
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Kathren and Moore 1986; Polednak and Frome 1981; Reyes et al. 1984; USNRC 1986). Clinical signs in one man following acute exposure to uranium did include dizziness and anorexia 6 days after exposure for 5 minutes to uranium tetrafluoride by inhalation (Zhao and Zhao 1990), and may have been related to rapidly developing renal disease. No etiology could be determined for increased central and peripheral nervous system diseases found in workers in a nuclear fuels fabrication plant (Hadjimichael et al. 1983). A series of studies by Gilman et al. (1999a, 1998b, 1998c) reported no brain lesions associated with ingestion of uranium in the drinking water. However, in other high-dose animal studies, neurological signs were reported in dogs, cats, rats, and guinea pigs. These signs included instability of gait indicative of neurological dysfunction in dogs and cats (Dygert 1949a); severe muscle weakness and lassitude from inhalation exposures in dogs and cats (Rothstein 1949a); central cholinergic neurological symptoms (piloerection, tremors, hypothermia, pupillary size decrease, exophthalmos) in rats from oral exposures (Domingo et al. 1987); and irritability, hyperactivity, upset equilibrium, rigidity of limbs, and respiratory arrest in rabbits from 4-hour dermal exposures (Orcutt 1949). However, no neurological effects were observed in rabbits orally exposed to 20 times larger oral doses of the same compound for 91 days. In view of the findings of the human and animal studies, it is doubtful that human exposure to uranium compounds at or near hazardous waste sites could result in damage to the nervous system.
Recent studies suggest that intramuscular deposition of uranium metal may result in neurological effects. Implantation of depleted uranium pellets in rats resulted in measurable uranium in the brain at 6–18 months after implantation (Pellmar et al. 1999a) and was accompanied by electrophysiological changes in hippocampal slices from the treated animals at 6 months (Pellmar et al. 1999b). In addition, military veterans with retained depleted uranium shrapnel fragments had lowered performance scores on computerized tests assessing performance efficiency which correlated with their urinary uranium levels (McDiarmid et al. 1999a). The etiology of these effects is unclear, although central nervous system toxicity due to other heavy metals (e.g., lead, mercury) is well documented. Further research is needed to confirm these results and determine the relevance of the effects from this unique exposure pathway to inhalation, oral, and dermal exposure.
Reproductive Effects. The existing human data from studies of uranium miners, millers, and processors (Muller et al. 1967; Waxweiler et al. 1981b; Wiese 1981) and most data from animal studies (Gilman et al. 1998a, 1998b; Leach et al. 1970; Llobet et al. 1991; Paternain et al. 1989) do not associate reproductive effects with uranium exposure.. Relatively high doses of uranium compounds (which also
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produced significant mortality in some cases) have resulted in some reproductive abnormalities manifested as significantly reduced sperm counts (Lloblet et al. 1991), reduced implantations and increased fetal resorptions and dead fetuses, maternal (reduced weight gain and food consumption, increased relative liver weight) and fetal toxicity (Domingo et al. 1989a); testicular lesions and degeneration and decreased testes weight (at near-lethal doses for long periods) (Malenchenko et al. 1978; Maynard et al. 1953); and reduced litter size (at a dose that produced 16% mortality) (Maynard et al. 1953) following intake of uranium compounds. However, no reproductive effects were found in a series of 28-day and 91-day uranium drinking water studies in rats and rabbits (Gilman et al. 1998a, 1998b, 1998c; Paternain et al. 1989).
In view of the lack of findings of reproductive effects in uranium miners, millers, and processors in numerous human studies and the equivocal findings in high-dose animal studies, it is doubtful that human exposure to uranium compounds at or near hazardous waste sites could result in interference with normal reproduction.
Developmental Effects. The present theories on the susceptibility of cells (with a high mitotic index such as are found in the embryo, fetus, and neonate) to damage by the DNA-adducting chemical action of uranium (as a heavy metal) (Cooper et al. 1982; Dungworth 1989; Stokinger 1981; Wedeen 1992) and ionization by high-LET, high specific-activity radiation (BEIR 1980, 1988, 1990; Muller et al. 1967; Otake and Schull 1984; Sanders 1986; Stokinger et al. 1953; UNSCEAR 1982, 1986, 1988) suggest that uranium may potentially interfere with normal development and may be teratogenic. However, no studies were located regarding the chemical or radiological effects of uranium on development in humans or animals following inhalation or dermal exposure for any duration.
Evidence for potential developmental toxicity is provided by results from oral animal studies in which the following effects were reported for uranyl acetate: increased fetal mortality, reduced survivability, and reduced growth (Paternain et al. 1989); decreased litter size, decreased viability and lactation indices, and pup liver weights (Domingo et al. 1989b); reduced fetal body weight and length, an increased incidence of stunted fetuses, increases in external and skeletal malformations and developmental variations, an increased incidence of cleft palate, underdeveloped renal papillae, and bipartite sternebrae, reduced or delayed ossification of the hind limb, fore limb, skull, and tail, an increase in the relative brain weight of the offspring, and reduced viability and lactation index (Domingo et al. 1989a); and embryotoxicity (Paternain et al. 1989). These effects were seen at relatively high doses far above any plausible human exposure.
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The investigators of a study in which relative liver weights were significantly decreased at all exposure levels (p<0.05 in the 0.028–0.28 mg U/kg/day group and p6 mg U/kg/day in a soluble form as uranyl acetate. At doses of 14 mg U/kg/day but not 6 mg U/kg/day, embryolethality (increased total and late resorptions, decreased number of live fetuses) was observed on gestation day 13 in dams exposed from 14 days prior to mating through gestation (Paternain et al. 1989). Gavage exposure over gestation days 6–15 resulted in an increased incidence of skeletal abnormalities (bipartite sternebrae, reduced and/or delayed ossification) at 14 and 28 mg U/kg/day and cleft palate at 6, 14, and 28 mg U/kg/day (Domingo et al. 1989a). Underdeveloped renal papillae were also observed but were not dose-related. Exposure of dams from late pregnancy (gestation day 13) continuing throughout lactation (21 days postpartum) resulted in reduced pup viability at 28 mg U/kg/day, but not at lower doses (Domingo et al. 1989b). Postpartum
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developmental events (pinna attachment, eye opening, incisor eruption) were unaffected at all doses. While developmental toxicity can be produced in animal models, the doses required are relatively high compared to known human exposures and are similar to a dose of 14.3 mg U/kg of uranyl nitrate that produced nausea, vomiting, and diarrhea in one human (Butterworth 1955).
Information on the pharmacokinetics of uranium in children is very limited. Since the skeletons of children are growing (higher rate of bone formation), it is possible that a higher fraction of circulating uranium will be deposited in bone than in adults. A study of uranium content in bone from three age groups (100 μg/L is indicative of recent absorption, while a concentration of 90% enriched) is used in special research reactors (most of which have been removed from operation) (Weigel 1983), nuclear submarine reactor cores, and nuclear weapons. Depleted uranium metal (DU) is used as radiation shielding, missile projectiles, target elements in plutonium production reactors, a gyroscope component, and counterweights or stabilizers in aircraft.
Uranium continuously undergoes transformation through the decay process whereby it releases energy to ultimately become a stable or nonradioactive element. For the uranium isotopes, this is a complex process involving the serial production of a chain of decay products, called progeny, until a final stable element is formed. The decay products of the uranium isotopes, which are also radioactive, are shown in Table 3-4. 238U is the parent isotope of the uranium series (234U is a decay product of 238U), while 235U is the parent isotope of the actinide series. All natural uranium isotopes and some of their progeny decay by emission of alpha particles; the other members of both series decay by emission of beta particles and gamma rays (Cowart and Burnett 1994). Both the uranium and the actinide decay series have three features in common. Each series begins with a long-lived parent, 235U or 238U, each series contains an isotope of the noble gas radon, and each series ends with a stable isotope of lead, 207Pb or 206Pb.
The amount of time required for one-half of the atoms of a radionuclide to transform is called its radioactive half-life. The rate of decay, and thus the half-life, for each radionuclide is unique. The half-life of 238U is very long, 4.5¥109 years; the half-lives of 235U and 234U are orders of magnitude lower, 7.1¥108 years and 2.5¥105 years, respectively. Since the activity of a given mass of uranium depends on the mass and half-life of each isotope present, the greater the relative abundance of the more rapidly decaying 234U and 235U, the higher the activity will be (EPA 1991). Thus, depleted uranium is less radioactive than natural uranium and enriched uranium is more radioactive.
Uranium is unusual among the elements because it is both a chemical and a radioactive material. The hazards associated with uranium are dependent upon uranium’s chemical and physical form, route of intake,
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and level of enrichment. The chemical form of uranium determines its solubility and, thus, transportability in body fluids as well as retention in the body and various organs. Uranium’s chemical toxicity is the principal health concern, because soluble uranium compounds cause heavy metal damage to renal tissue. The radiological hazards of uranium may be a primary concern when inhaled, enriched (>90%) and insoluble uranium compounds are retained long-term in the lungs and associated lymphatics.
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4.1 PRODUCTION
Uranium is present in the earth’s crust at approximately 3 ppm (2 pCi/g) (du Preez 1989). Although there are more than 100 uranium ores, carnotite, pitchblende, coffinite, uraninite, tobernite, autunite, tyuyamunite, and a few others are the main ores of commercial interest. The main ores are described in Table 4-1. The most economically attractive uranium ores have uranium concentrations in excess of 1,000 ppm (700 pCi/g) (Stokinger 1981; Weigel 1983). In the United States, the major ore deposits are located in Colorado, Utah, Arizona, New Mexico, Wyoming, Nebraska, and Texas (EPA 1985a). The steps necessary to produce uranium for its various uses include mining, milling, conversion to uranium hexafluoride (UF6), enrichment, reduction to metal or oxidation to uranium oxide, and fabrication into the desired shape. The steps for preparing commercial reactor grade, submarine reactor grade, or weapons-grade uranium are the same, except the last two require a more aggressive enrichment process. Depleted uranium metal is produced by reducing the depleted uranium hexafluoride byproduct. Conventional fabrication methods are used to configure the uranium for specific uses, such as rectangular solid blocks for helicopter rotor counterbalances and parabolic or cylindrical solids for military depleted uranium projectiles.
Mining. Open-pit mining, in situ leaching, and underground mining are three techniques that have been used for mining uranium-containing ores (EPA 1985a). Uranium is found in all soil and rock, but the higher concentrations found in phosphate rock, lignite, and monazite sands are sufficient in some areas for commercial extraction (Lide 1994). The two most commonly used mining methods are open-pit and underground mining. The choice of method is influenced by factors such as the size, shape, grade, depth, and thickness of the ore deposits (Grey 1993). In situ leaching involves leaching (or dissolving) uranium from the host rock with liquids without removing the rock from the ground and can only be carried out on unconsolidated sandstone uranium deposits located below the water table in a confined aquifer. A leaching solution is introduced into or below the deposit and pumped to the surface, where the uranium-pregnant liquor is processed in a conventional mill to precipitate the uranium as yellowcake (U3O8 and other oxides) (DOE 1995b; Grey 1993).
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Milling. Ore mined in an open-pit or underground mine is crushed and leached in a uranium mill. The initial step in conventional milling involves crushing, grinding, and wet and/or dry classification of the crude ore to produce uniformly sized particles that are similar in size to beach sand. A slurry generated in the grinding circuit is transferred to a series of tanks for leaching by either an alkaline or acid process. Generally, leaching is a simple process whereby uranyl ions are extracted by a solvent. Uranyl ions are stripped from the extraction solvent and precipitated as yellowcake, predominantly U3O8 (EPA 1995d). Yellowcake is pressed, dried, banded, and shipped for refinement and enrichment. Some of the process streams can also be used to extract other oxides, such as vanadium pentoxide. The byproduct of this process is the leftover sand, known as tailings. Thus, tailings are the original sand minus much of the uranium plus residual process chemicals and tailings are less radioactive than the original ore. (Uranium metal production in a conversion facility is done post-enrichment.) Generalized flow charts for the alkaline and acid leaching processes for ore concentration and uranium production are shown in Figure 4-1.
Enrichment. Next, the U3O8 is chemically converted to UF6. The enrichment process increases, or enriches, the percentage of the fissionable 235U isotope, as well as 234U. In the United States, the process used for enrichment is gaseous diffusion. The mechanism for enrichment is based on the fact that a UF6 molecule containing 235U or 234U is lighter and smaller, and has, therefore, a slightly higher thermal velocity than a UF6 molecule containing 238U. As the UF6 passes through the series of diffusion stages, the 234UF6 and 235UF6 molecules gradually become more concentrated downstream and less concentrated upstream, while the 238UF6 concentrates conversely. The lead portion of the stream is collected and recycled to reach the desired enrichment. The tail portion containing a reduced 235UF6 content called depleted UF6 can be stored in the vicinity of the gaseous diffusion plant sites (DOE 1994b). There are an estimated 560,000 metric tons of depleted uranium currently in storage as UF6. A second enrichment technology, gas centrifuge separation, has been used in Europe. A third technology, laser separation, is currently under development (DOE 1995b). A fourth technology, thermal separation, is inefficient and no longer used.
Fuel fabrication. The enriched UF6 is either reduced to metallic uranium and machined to the appropriate shape, or oxidized to uranium dioxide and formed into pellets of ceramic uranium dioxide (UO2). The pellets are then stacked and sealed inside metal tubes that are mounted into special fuel assemblies ready for use in a nuclear reactor (DOE 1995b; Uranium Institute 1996).
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4. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL
Product fabrication. Uranium metal has commercial and industrial uses due to its great density and strength. It is alloyed with a range of metals to meet other commercial and industrial needs. As with steel, uranium can be formed and fashioned by drop forging, dye casting, and machining and is often painted to minimize oxidation. Some well known uses for these products are gyroscopic wheels in guidance systems, helicopter rotor blade counterbalances, weights in airplane control surfaces, and radiation shields for high radioactivity sources (e.g. industrial radiography).
Production. Uranium production from 1975 to 1996 is shown in Table 4-2. Peak production of uranium occurred in 1980 at 21,852 short tons (1.98×107 kg) and decreased until 1993. This was the same period when the planning and construction of new nuclear power plants ceased in the United States. Production of U3O8 had decreased to 4,443 short tons (4.03×106 kg) in 1990 and to 1,534 short tons (1.39×106 kg) in 1993, a 65% reduction (ABMS 1994; EPA 1985a). In 1996, U.S. uranium production was 3,160 (2.87×106 kg) short tons, an increase of 5% from the 1995 level and the highest level since 1991 (DOE 1996a). Underground and open-pit mining have been the two most commonly used methods of mining uranium ores. However, by 1994, uranium was produced primarily by in situ leaching methods. A summary of U.S. mine production from 1985 through 1996 (see Table 4-3) illustrates the shift from underground and open-pit mining to in situ leaching.
Leached uranium concentrate was produced in 1996 in Wyoming, Louisiana, Nebraska, New Mexico, and Texas. At the end of 1996, two phosphate by-product plants and five in situ leaching plants were in operation. In addition, seven phosphate by-product and in situ leaching plants were inactive, and seven conventional uranium mills were being maintained in stand-by mode (DOE 1996b).
4.2 IMPORT/EXPORT
The importation of uranium increased significantly in the 1980s (EPA 1985a). In 1983, 3,960 short tons of U3O8 equivalent were imported into the United States (USDOC 1984), which was about 37% of the domestic production. In 1987, the amount of U3O8 equivalent imported into the United States was 5,630 short tons (USDOC 1988). The amounts of uranium and uranium compounds imported into the United States during the period 1989–1997 are presented in Table 4-4 (USDOC 1995, 1997). The importation of uranium and uranium compounds peaked in 1990 at about 23 million kg (about 1 million tons) and has remained approximately the same, with some fluctuation, since that time.
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The amount of uranium and uranium compounds exported from the United States during 1989–1993 is shown in Table 4-5. The total volume of uranium and uranium compounds exported during 1989–1993 was two orders of magnitude lower than the quantities imported during this same time period. Exports in 1996 were 5.2 million kg. Most of the foreign sales (Canada, France, Germany, Japan, South Korea, United Kingdom) occurred after the uranium entered the U.S. market as imports (DOE 1999b).
4.3 USE
Uranium has been produced for use in the commercial nuclear power industry as low-enriched metal or ceramic UO2 fuel pellets; smaller quantities of high-enriched fuel are produced for U.S. Navy ships and for weapons manufacture (EPA 1985b; Stokinger 1981). Uranium fuel lasts months to years before refueling is needed, and then only a small fraction of the uranium has actually been fissioned, making fuel reprocessing an option used in other countries. One pound of completely fissioned uranium produces the same amount of energy as 1,500 tons of coal (Lide 1994). Depleted uranium is used in the manufacture of armor-piercing ammunition for the military, in inertial guidance devices and gyro compasses, as counterbalances for helicopter rotors, as counterweights for aircraft control surfaces, as radiation shielding material, and as x ray targets (EPA 1985b; USDI 1980). Uranium dioxide is used to extend the lives of filaments in large incandescent lamps used in photography and motion picture projectors. Uranium compounds are used in photography for toning, in the leather and wood industries for stains and dyes, and in the silk and wood industries as mordants. Ammonium diuranate is used to produce colored glazes in ceramics. Uranium carbide is a good catalyst for the production of synthetic ammonia (Hawley 1981). Additionally, uranium was used in dental porcelains for many years, but this practice has been discontinued (Thompson 1976). According to the USDI (1980), the major uses of depleted uranium in 1978 were military ammunition, 71.8%; counterweights, 11.4%; radiation shielding, 13.6%; and chemical catalysts, 3.2% although this ratio may shift to support war efforts.
4.4 DISPOSAL
Radioactive waste containing uranium is usually grouped into three categories: uranium mill tailings, low-level waste, and, in the case of spent reactor fuel, high-level waste.
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Uranium mill tailings are the residual sand and trace chemicals from the processing of uranium ore. About 150 tons of enriched uranium are required per year to fuel a 1,000-megawatt electric nuclear power reactor, and about 500 times that amount of ore is required to obtain the uranium. The total accumulation of uranium mill tailings in the United States is approximately 140 million tons (Murray 1994). Tailings from mines and mills that process other metals should also be expected to contain elevated concentrations of uranium and its progeny, although this may not be readily recognized.
Disposal methods for processed uranium tailings have been discussed by Bearman (1979). In the late 1940s, mainly unconfined disposal systems were used. Untreated solid wastes were stored as open piles and, in some cases, were spread in urban areas where they were used as fill and as the sand in concrete used to build roads, walks, drives, and concrete block, and in brick mortar. As a result of the Animas River Survey in the United States, tailing control programs were instituted in 1959 to prevent airborne and waterborne dispersal of the wastes. Confined disposal methods were devised to reduce the exposure and dispersion of wastes and to reduce seepage of toxic materials into groundwater to the maximum extent reasonably achievable. Under the Uranium Mill Tailings Radiation Control Act (UMTRCA) of 1978, the
U.S. Department of Energy (DOE) designated 24 inactive tailings piles for cleanup. These 24 sites contained a total of about 28 million tons of tailings and covered a total of approximately 1,000 acres (EPA 1985b). Cleanup has been completed at some sites.
In 1977, the EPA issued Environmental Radiation Protection Standards to limit the total individual radiation dose due to emissions from uranium fuel cycle facilities, including licensed uranium mills. This standard specified that the “annual dose equivalent does not exceed 25 millirems (0.25 mSv) to the whole body, 75 millirems (0.75 mSv) to the thyroid, and 25 millirems (0.25 mSv) to any other organ of any member of the public as the result of exposures of planned discharges of radioactive materials…to the general environment” (40 CFR 190). The EPA also established environmental standards for cleanup of open lands and buildings contaminated with residual radioactivity from inactive uranium processing sites (40 CFR 192).
Low-level radioactive waste (LLRW), which may contain uranium, is disposed of at DOE facilities and at commercial disposal facilities. Since 1963, six commercial LLRW facilities have operated, but only two were in operation in 1995. A 1992 report listed the total volume of LLRW buried at all 6 sites to be approximately 50 million cubic feet (Murray 1994). Only a small fraction of the LLRW contains uranium.
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The method of disposal for commercial and DOE LLRW has been shallow land burial, in which the waste is disposed of in large trenches and covered. This method of disposal relies upon natural features to isolate the waste. Although U.S. Nuclear Regulatory Commission (USNRC) regulations for LLRW disposal (10 CFR 61) permit shallow land burial, many states have enacted more stringent regulations that require artificial containment of the waste in addition to natural containment (Murray 1994). The EPA has proposed regulations for LLRW disposal that would apply to DOE facilities (EPA 1998b).
High-level radioactive waste (HLRW) includes spent fuel, which is the uranium fuel rods that have been used in a nuclear reactor. When the fuel rods are removed from the reactor for refueling, they still contain most of the original unfissioned uranium. However, the hazard from the large activity of fission products and plutonium that have been produced in the fuel rods overshadows that of uranium. Approximately 30,000 metric tons of spent fuel have been removed from U.S. power reactors through 1994 (Murray 1994). There is currently no permanent disposal facility for HLRW in the United States; these wastes are being stored at commercial nuclear power plants and DOE facilities where they were produced. The NRC has issued standards for the disposal of HLRW (10 CFR 60), and the DOE is pursuing the establishment of an HLRW facility. Efforts to establish an HLRW facility, which began over two decades ago, have experienced many delays. A facility for the permanent disposal of HLRW is not projected to be in operation before 2010 (Murray 1994).
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5.1 OVERVIEW
Elevated levels of uranium have been identified in at least 54 of the 1,517 current or former EPA National Priorities List (NPL) hazardous waste sites (HazDat 1999). However, the number of sites evaluated for uranium is not known. The distribution of these sites within the United States is shown in Figure 5-1.
Uranium is a naturally occurring radioactive element that is present in nearly all rocks and soils; it has an average concentration in U.S. soils of about 2 pCi/g (3 ppm) (du Preez 1989; NCRP 1984a). Some parts of the United States, particularly the western portion, exhibit higher than average uranium levels due to natural geological formations. Most uranium ores contain between 0.05 and 0.2% uranium, up to 1,000 times the levels normally found in soil (Uranium Institute 1996).
Uranium can undergo oxidation-reduction reactions in the environment or microbial reactions to form complexes with organic matter (Premuzie et al. 1995). The only mechanism for decreasing the radioactivity of uranium is radioactive decay. Since all three of the naturally occurring uranium isotopes have very long half-lives (234U, 2.4×105 years; 235U, 7.0×108 years; and 238U, 4.5×109 years), the rate at which the radioactivity diminishes is very slow (NCRP 1984a). Therefore, the activity of uranium remains essentially unchanged over periods of thousands of years.
Uranium may be redistributed in the environment by both anthropogenic and natural processes. The three primary industrial processes that cause this redistribution are operations associated with the nuclear fuel cycle that include the mining, milling, and processing of uranium ores or uranium end products; the production of phosphate fertilizers for which the phosphorus is extracted from phosphate rocks containing uranium; and the improper disposal of uranium mine tailings (Cottrell et al. 1981; Hart et al. 1986; NCRP 1984a; Yang and Edwards 1984). Essentially no uranium is released from nuclear power plants because of the fuel assembly design and the chemical and physical nature of the uranium oxide fuel. Examples of uranium redistribution by natural processes include activities and processes that move soil and rock, such as resuspension of soils containing uranium through wind and water erosion, volcanic eruptions (Kuroda et al. 1984), operation of coal-burning power plants (coal containing significant quantities of uranium), and construction of roads and dams.
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Uranium becomes airborne due to direct releases into the air from these processes. Deposition of atmospheric uranium may occur by wet (rain, sleet, or snow) or dry (gravitation or wind turbulence) processes. The rate of uranium deposition is dependent upon such factors as particle size, particle density, particle concentration, wind turbulence, and chemical form. Data are lacking on residence times of particulate uranium in the atmosphere, although UNSCEAR (1988) assumed that it behaves like atmospheric dust, for which meteorological models exist.
Uranium deposited by wet or dry precipitation will be deposited on land or in surface waters. If land deposition occurs, the uranium can be reincorporated into soil, resuspended in the atmosphere (typically factors are around 10-6), washed from the land into surface water, incorporated into groundwater, or deposited on or adsorbed onto plant roots (little or none enters the plant through leaves or roots). Conditions that increase the rate of formation of soluble complexes and decrease the rate of sorption of labile uranium in soil and sediment enhance the mobility of uranium. Significant reactions of uranium in soil are formation of complexes with anions and ligands (e.g., CO3-2, OH-1) or humic acid, and reduction of U+6 to U+4. Other factors that control the mobility of uranium in soil are the oxidation-reduction potential, the pH, and the sorbing characteristics of the sediments and soils (Allard et al. 1979, 1982; Brunskill and Wilkinson 1987; Herczeg et al. 1988; Premuzie et al. 1995).
Uranium in surface water can disperse over large distances to ponds, rivers, and oceans. The transport and dispersion of uranium in surface water and groundwater are affected by adsorption and desorption of uranium on aquatic sediments. As with soil, factors that control mobility of uranium in water include oxidation-reduction potential, pH, and sorbing characteristics of sediments and the suspended solids in the water (Brunskill and Wilkinson 1987; Swanson 1985). In one study of a stream with low concentrations of inorganics, low pH, and high concentrations of dissolved organic matter, the concentration of uranium in sediments and suspended solids was several orders of magnitude higher than in the surrounding water because the uranium was adsorbed onto the surface of the sediments and suspended particles (Brunskill and Wilkinson 1987).
The levels of uranium in aquatic organisms decline with each successive trophic level because of very low assimilation efficiencies in higher trophic animals. Bioconcentration factors measured in fish were low (Mahon 1982; Poston 1982; Waite et al. 1988) and were thought to arise from the extraction of uranium from the water or simply from the accumulation of uranium on gill surfaces (Ahsanullah and Williams
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1989). In plants, uptake of uranium may be restricted to the root system and may actually represent adsorption to the outer root membrane rather than incorporation into the interior of the root system (Sheppard et al. 1983). Most of this uranium may be removed by washing the vegetable surfaces; cutting away the outer membrane will essentially result in complete removal. No significant translocation of uranium from soil to the aboveground parts of plants has been observed (Van Netten and Morley 1983).
The EPA has established a nationwide network called the Environmental Radiation Ambient Monitoring System (ERAMS) for obtaining data concerning radionuclides, including natural uranium isotopes, in environmental media. Sampling locations for ERAMS were selected to provide optimal population coverage (i.e., located near population centers). Airborne uranium concentrations and precipitation levels of uranium were quite low, in the attocurie/m3 (10-3 nanoBq/m3) and 0.006–0.098 picocurie/L (0.0002–0.004 Bq/L/m3) ranges, respectively (EPA 1994). However, both air samples and water samples taken near facilities producing uranium ore or processing uranium were found to be higher, in the pCi/L range (Eadie et al. 1979; Lapham et al. 1989; Laul 1994; NCRP 1984a; Tracy and Meyerhof 1987). The ERAMS reports document 234U to 238U concentration ratios in drinking water which deviate from the equilibrium value of unity found in undisturbed crustal rock. Theories proposed to account for this natural phenomenon involve water contact with soil and permeable rock containing uranium. The 238U atoms transform through 234Th to 234U, and any process which removes either of these radionuclides from the solid changes the 234U to 238U ratio. 238U atoms at the solid-liquid interface which emit decay alpha particles inward may experience a kinetic energy recoil sufficient to either tear the 234Th progeny from the solid or fracture the surficial solid layer making the 234Th more accessible for the enhanced dissolution that thorium typically experiences relative to uranium in mineral matrices. Either process can enhance the relative 234U content of the liquid. Should that liquid stabilize in another location and evaporate, the localized solids could show an enhanced 234U ratio.
A large drinking water study was performed in which data from the National Uranium Resource Evaluation (NURE) program plus data prepared for the EPA (Drury 1981) were compiled for a total of over 90,000 water samples. Domestic water supplies were represented by 28,000 samples and averaged
1.7 pCi/L (2.5 μg/L) uranium, with a population weighted mean value for finished waters, based on 100 measurements, of 0.8 pCi/L (1.2 μg/L). Other studies show the population weighted average concentration of uranium in U.S. community drinking water to range from 0.3 to 2.0 pCi/L (0.4–3.0 μg/L)
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(Ohanian 1989), while concentrations of uranium from selected drinking water supplies analyzed by EPA laboratories were generally 50%) so that wastes from uranium milling contain only low levels of uranium; however, the levels of uranium progeny (e.g., radium) remain essentially unchanged. Uncontrolled erosion of these wastes from open tailings piles not protected from the weather occurred at a Shiprock, New Mexico, uranium mill site, resulting in contamination of the surrounding area (Hans et al. 1979). Uncontrolled erosion also occurred in storage areas such as the St. Louis Airport Storage Site in Missouri (Seelley and Kelmers 1985). Increased levels of uranium, radium, radon, and other decay products of uranium have also been measured around these sites, particularly in the soil. A number of controlled disposal locations on government-owned mill sites exist, but the ones identified involved uncontrolled disposal.
At various facilities that process uranium for defense programs, uranium is released to the atmosphere under controlled conditions, resulting in deposition on the soil and surface waters. Monitoring data from the area surrounding the Fernald Environmental Management Project (formerly the Fernald Feed Materials Production Center) showed that soil contained uranium released from the facility (Stevenson and Hardy 1993).
The uranium content of phosphate rock, a source of phosphorus for fertilizers and phosphoric acid for the chemical industry, ranges from several pCi/g to 130 pCi/g (several ~200 μg/g) (Boothe 1977; UNSCEAR 1977, 1982). During milling, much of the uranium content becomes concentrated in slag by-products (Melville et al. 1981). The slag by-products are often used for bedrock in the paving of roads, thus transferring the uranium-rich slag to the soil (Melville et al. 1981; Williams and Berven 1986). Because of the large amounts of phosphate fertilizer produced annually (12–15 million tons), trace amounts of uranium progeny remaining in the fertilizer result in the distribution of about 120 Ci (180 metric tons) per year over
U.S. agricultural lands (Kathren 1984).
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Combustion of coal is a significant source of enhanced natural radioactivity (especially combustion of coal from the western United States, which contains significantly more uranium than coal from the eastern United States). When coal is burned, some of the radioactivity is released directly to the atmosphere, but a significant fraction is retained in the bottom ash. Enhanced concentrations of uranium have been found on the ground around coal-fired power plants (UNSCEAR 1982).
Unauthorized landfill disposal of uranium processing wastes (e.g., Shpack Landfill in Norton, Massachusetts, and the Middlesex Municipal Landfill in Middlesex, New Jersey) has resulted in soil contamination (Bechtel National 1984; Cottrell et al. 1981). Also, elevated uranium concentrations have been measured in soil samples collected at 30 of 51 hazardous waste sites and in sediment samples at 16 of 51 hazardous waste sites (HazDat 1998). The HazDat data includes both Superfund and NPL sites. Elevated concentrations of uranium have been detected in soil, in surface water, in groundwater, or in all three of these environmental media from these sites. In several cases, the uranium concentrations in soils were significantly elevated. For example, uranium concentrations from the Shpack/ALI site were found to be 16,460 pCi/g (24,000 μg/g). At the United States Radium Corporation site (New Jersey), uranium concentrations ranged from 90 to 12,000 pCi/g (130–18,000 μg/g); for the Monticello site (Utah), uranium levels were reported to range from 1 to 24,000 pCi/g (1.5–36,000 μg/g) (HazDat 1998).
5.3 ENVIRONMENTAL FATE
5.3.1 Transport and Partitioning
The components of an ecosystem can be divided into several major compartments (Figure 5-3) (NCRP 1984b). None of the environmental compartments exist as separate entities; they have functional connections or interchanges between them. Figure 5-3 also shows the transport pathways between the released uranium and the environmental compartments as well as the mechanisms that lead to intakes by the population. Initial uranium deposition in a compartment, as well as exchanges between compartments (mobility), are dependent upon numerous factors such as chemical and physical form of the uranium, environmental media, organic material present, oxidation-reduction potential, nature of sorbing materials, and size and composition of sorbing particles.
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Natural processes of wind and water erosion, dissolution, precipitation, and volcanic action acting on natural uranium in rock and soil redistribute far more uranium in the total environment than the industries in the nuclear fuel cycle. However, those industries may release large quantities of uranium in specific locations, mainly in the form of solids placed on tailings piles, followed by liquids released to tailings ponds and then airborne releases, both directly from the facilities and by wind erosion of the tailings piles. Although solid releases represent the largest quantity of uranium redistribution, they remain on the facility grounds and are normally inaccessible to the public. It is the airborne (direct and wind erosion on tailings piles) and liquid releases (tailings pond runoff and water erosion of tailings) which most likely represent the important pathways for public exposure (i.e., inhalation and ingestion) if pathways can be completed.
While entrained in the air, particulate uranium represents an inhalation source for humans, the extent of which is dependent upon concentration and particle size. For particulate uranium to be an inhalation hazard to humans, the particulates must be in the size range of 1–10 μm (Bigu and Duport 1992; ICRP 1979). In some cases, the solid tailings have been removed from the site for use as fill or construction material, which can lead to external radiation exposures primarily from the uranium progeny.
Deposition of the atmospheric uranium can occur by dry deposition or wet deposition (Essien et al. 1985). Dry deposition results from gravitational settling and impaction on surfaces exposed to turbulent atmospheric flow. The rate of dry deposition is dependent upon particle size distribution, chemical form, particle density, and degree of air turbulence. Few experimental data on the particle size and residence time of uranium and uranium compounds present in ambient atmospheres are available; however, uranium particles are expected to behave like other particles for which data are available, which show that smaller uranium particles (<5 μm) travel longer distances than larger particles because of their longer residence time in the atmosphere due to their low settling velocity.
The chemical form of the uranium affects the atmospheric residence time. One uranium compound for which there are data regarding residence time and particle size is uranium hexafluoride, a soluble compound, which will hydrolyze in the atmosphere to particulate UO2F2•nH2O and hydrogen fluoride gas (Bostick et al. 1985). In the case of UO2F2, although the particles were small (<2.5 μm), its atmospheric residence time was estimated to be only 35 minutes as a result of rapid hydration and agglomeration to larger particles that have faster settling velocities (Bostick et al. 1985).
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In wet deposition of airborne contaminants, the uranium is washed from the atmosphere by rain, sleet, snow or other forms of moisture. The rate of wet deposition depends upon particle size and solubility (chemical form).
Uranium thus deposited (dry or wet) will usually reside on land or be deposited on surface waters. If land deposition occurs, the uranium can incorporate into the soil or adhere to plant surfaces, be resuspended in the atmosphere as a result of wind action, or be washed from the land into surface water and groundwater. Resuspension factors are typically quite low (10-6) and protective against significant exposures, but this may not apply to windy and arid areas. Resuspension into the air can be an inhalation source even after the plume or source has disappeared.
In addition to the migration of dissolved or suspended uranium due to the movement of water in the environment, the transport and dispersion of uranium in surface water and groundwater are affected by adsorption and desorption of the uranium on surface water sediments. On the other hand, migration of uranium in soil and subsoil and uptake in vegetation are usually quite local involving distances from several centimeters to several meters.
In most waters sediments act as a sink for uranium and the uranium concentrations in sediments and suspended solids are several orders of magnitude higher than in surrounding water (Brunskill and Wilkinson 1987; Swanson 1985). Factors that control the mobility of uranium from sediment to the water phase are the oxidation-reduction potential, the pH, the characteristics of complexing agents or ligands, and the nature of sorbing materials in the water. Inorganic or organic ligands that can form soluble complexes with uranium will result in mobilization of the uranium in water. However, the stability of such complexes is dependent on the pH. For example, uranium is likely to be in solution as a carbonate complex in oxygenated water with high alkalinity (Herczeg et al. 1988); however, in acidic waters (pH <6 containing low concentrations of inorganic ions and high concentrations of dissolved organic matter), the uranium is in solution as the soluble organic complex (Brunskill and Wilkinson 1987).
The oxidation-reduction potential of water is important in controlling the mobility of uranium. In anoxic waters where the aquatic environment is reductive, U(VI) will be reduced to U(IV) (e.g., changed from a soluble compound to an insoluble one). The U(IV) will be deposited into the sediment due to the insolubility of the resulting U(IV) salts (Allard et al. 1979; Herczeg et al. 1988). Mobilization and
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deposition of uranium as defined by the oxidation-reduction potential of the water has been observed by several investigators (Barnes and Cochran 1993; Shaw et al. 1994). Uranium can also be removed from solution by physical adsorption processes, such as adsorption onto oxides of iron or manganese that occur as coatings on the particles of soil and sediment (Ames et al. 1982).
The mobility of uranium in soil and its vertical transport (leaching) to groundwater depend on properties of the soil such as pH, oxidation-reduction potential, concentration of complexing anions, porosity of the soil, soil particle size, and sorption properties, as well as the amount of water available (Allard et al. 1982; Bibler and Marson 1992). Retention of uranium by the soil may be due to adsorption, chemisorption, ion exchange, or a combination of mechanisms (Allard et al. 1982). Any soil property that alters the sorption mechanism will also alter the mobility of uranium in the soil. The sorption of uranium in most soils is such that it may not leach readily from soil surface to groundwater, particularly in soils containing clay and iron oxide (Sheppard et al. 1987), although other geological materials such as silica, shale, and granite have poor sorption characteristics (Bibler and Marson 1992; Erdal et al. 1979; Silva et al. 1979; Tichnor 1994).
Sorption in most soils attains a maximum when the neutral hydroxy complex of uranium is at a maximum. However, at pH 6 and above, and in the presence of high carbonate or hydroxide concentrations, uranium may form anionic complexes such as [UO2(OH)4]-2. The mobility of anionic uranium complexes in soil is dependent upon the nature of the soil. For example, the decrease in sorption in soil with little anion-exchange capacity may result in increased mobility; however, increased sorption in soil with high anion-exchange may result in decreased mobility (Allard et al. 1982; Ames et al. 1982; Brookins et al. 1993; Ho and Doern 1985; Hsi and Langmuir 1985; Tichnor 1994).
Other factors also affect the mobility of uranium in soil. A field study performed near an active carbonate leach uranium mill showed that uranium in an alkali matrix can migrate to the groundwater (Dreesen et al. 1982). Uranium mobility may also be increased due to the formation of soluble complexes with chelating agents produced by microorganisms in the soil (Premuzie et al. 1995).
Uranium may be transported to vegetation by air or by water. It can be deposited on the plants themselves by direct deposition or resuspension, or it can adhere to the outer membrane of the plant's root system with potential limited absorption. Similarly, uranium deposited on aquatic plants or water may be adsorbed or taken up from the water. The plants, aquatic or terrestrial, may be eaten directly by humans or consumed
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by land or aquatic animals, which provide food for humans. The uptake or bioconcentration of uranium by plants or animals is the mechanism by which uranium in soil, air, and water enters into the food chain of humans.
Numerous factors influence the bioaccumulation of uranium, such as the chemical and physical form of the uranium; the season of the year and other climatic factors such as temperature, age of the organism, specific tissue or organs involved; and the specific characteristics of the local ecosystem, such as total suspended and dissolved solids. Bioconcentration factors for uranium have been measured by several investigators in various aquatic organisms. Mahon (1982) measured bioconcentration factors of 1,576 and 459 in algae and plankton, respectively. Horikoshi et al. (1981) determined bioconcentration factors in several species of bacteria that ranged from 2,794 to 354,000. However, bioconcentration by the bacteria represented adsorption onto the cell surfaces of the bacteria rather than true biological uptake.
Low bioconcentration factors for uranium were observed in fish. The highest bioconcentration factors observed in fillet of rainbow trout (Salmo gairdneri), white and finescale suckers (Castastomus catactomus), and lake whitefish (C. clupeaformis) did not exceed a value of 38 (Mahon 1982; Poston 1982; Swanson 1983, 1985). Ahsanullah and Williams (1989) concluded that the primary source of uranium for crab (Pachygrapsus laevimanus) and zebra winkle (Austrocochlea constricta) was from water since both fed and starved animals took up uranium at the same rate.
Uranium is transported poorly from soils to plants (Dreesen et al. 1982; Moffett and Tellier 1977). As with aquatic organisms, the uptake of uranium by plants is dependent on the nature of the soils (soil texture and organic content), the pH, and the concentration of uranium in the soil. Greater plant uptake is expected to occur in soils that contain higher levels of available uranium (i.e., less sorption of uranium to soil particles or formation of soluble uranium complexes). Swiss chard grown in sandy soils contained 80 times the levels of uranium found in Swiss chard grown in peat soil (Sheppard et al. 1983). The uptake of uranium by native plants, expressed as plant/soil concentration ratio (CR), grown near a mining and milling complex was 0.8 compared to a CR of 0.09 for plants grown in soil with background levels of uranium (Ibrahim and Wicker 1988). The effect of soil and plant type on CR values has been reviewed by Mortvedt (1994).
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Reported CR values for plant/soil interaction vary widely (range, 0.0025–0.81) (Garten 1978; Ibrahim and Wicker 1988; Mortvedt 1994). Although some studies indicate that CR values in plants do not vary linearly with the concentration of uranium in the soil (Mortvedt 1994), other reported studies show a linear relationship between plant content and soil content of uranium (NCRP 1984a). It has been postulated that uranium uptake by plants may be limited to the outer membrane of the root system and may not occur on the interior of the root at all (Van Netten and Morley 1983; Sheppard et al. 1983). However, other investigators have reported the transfer of uranium from soil to the stems and leaves of plants in which the CR decreased in the following order: fruit < leaf < root (Morishima et al. 1977). Because of the higher root sorption of uranium, it has been postulated that consumption of radishes and other root vegetables grown in uranium-containing soils may be a source of human exposure (Van Netten and Morley 1983). Thorough cleansing of the plant exterior, especially if performed in conjunction with removal of the outer membrane, may remove most or all of the uranium.
5.3.2 Transformation and Degradation
5.3.2.1 Air
The presence of uranium and uranium compounds in the atmosphere results from activities associated with uranium mining, milling, processing, and use. There is limited information available regarding the abiotic transformation and degradation of uranium and uranium compounds, except for uranium hexafluoride. Uranium hexafluoride immediately hydrolyzes on contact with moisture in the air to form uranyl fluoride (UO2F2) and hydrofluoric acid (HF). Uranyl fluoride is hygroscopic and will absorb moisture from the air, resulting in an increased settling velocity associated with the larger particle size. The half-life of a release of airborne UF6 is about 35 minutes (Bostick et al. 1985). Uranyl fluoride is a stable oxohalide compound of uranium which is soluble in water, a factor that will increase its mobility in the environment once deposition from the air has occurred.
5.3.2.2 Water
The principal abiotic processes that transform uranium in water are formation of complexes and oxidation-reduction reactions that have been described in Section 5.3.1. In seawater at pH 8.2, it was shown that U(IV) exists as 100% neutral hydroxo complexes, and UO2+2 and U(VI) exist as 100% carbonato
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complexes. In freshwater at pH 6, U(IV) was shown to exist as 100% hydroxo complexes, and UO2+2 existed as 12% hydrated complexes, 18% hydroxo complexes, 8% fluoro complexes, and 60% carbonato complexes. In freshwater at pH 9, U(IV) exists as 100% hydroxo complexes, but UO2+2 exists as 100% carbonato complexes (Boniforti 1987).
Oxidation-reduction conditions are important in the geologic transport and deposition of uranium. Oxidized forms of uranium (U[VI]) are relatively soluble and can be leached from the rocks to migrate in the environment. When strong reducing conditions are encountered (e.g., presence of carbonaceous materials or H2S), precipitation of the soluble uranium will occur.
5.3.2.3 Sediment and Soil
The primary abiotic and biological processes that transform uranium in soil are oxidation-reduction reactions that convert U(VI) (soluble) to U(IV) (insoluble). Reduction of U(VI) to U(IV) can occur as a result of microbial action under anaerobic soil or sediment conditions, thereby reducing the mobility of uranium in its matrix (Barnes and Cochran 1993; Francis et al. 1989). Further abiotic and biological processes that can transform uranium in the environment are the reactions that form complexes with inorganic and organic ligands (see Section 5.3.1).
Certain microorganisms (e.g., Thiobacillus ferrooxidans) can facilitate the oxidation of Fe+2 to Fe+3. The Fe+3 ion, in turn, can convert insoluble uranium dioxide to soluble UO2+2 ions by the following reaction:
2Fe+3 + UO2 6 UO2+2 + 2Fe+2
This reaction enhances the mobility of uranium in soil from mining and milling wastes (Barnes and Cochran 1993; de Siloniz et al. 1991; Scharer and Ibbotson 1982).
Uranium may be removed from the pore water of sediments under sulfate reduction conditions; microbes may control this process indirectly (Barnes and Cochran 1993).
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5.4 LEVELS MONITORED OR ESTIMATED IN THE ENVIRONMENT
In 1973, the EPA established a nationwide network, called ERAMS, for obtaining data in environmental samples. ERAMS consists of a network of sampling stations that provide air, surface and drinking water, and milk samples that the EPA uses to obtain environmental concentrations of radioactive material. The objective of this system is to identify trends in the accumulation of long-lived radionuclides in the environment (EPA 1994). Sampling locations for ERAMS are located near primary population centers to provide optimal population coverage.
The ratio of 234U to 238U would be expected to be unity as long as the uranium stays locked inside undisturbed crustal rock in secular equilibrium with its progeny, but measurements show that the ratio is typically different than unity (EPA 1994). This disequilibrium occurs when the rock is disturbed by chemical or physical changes involving water. In the environment, a portion of the 234U separates from the 238U by what is theorized to be a physical process (alpha recoil ejection of the 232Th decay product from surfaces of soil particles) or a combination of physical and chemical processes (a 238U transformation at the soil particle surface fractures the surface allowing access for water to dissolve the more soluble 234Th product) (NCRP 1984a). These processes can change the uranium isotope ratios in air, soil, and water.
5.4.1 Air
For airborne particles collected for the ERAMS program, 234U, 235U, and 238U analyses are performed on semiannually composited air filters collected from continuously operating airborne particulate samplers. Following chemical separation, the uranium is quantified by α-spectroscopy.
Table 5-2 shows the results of monitoring for uranium in airborne particles for the January to June 1993 composites as published in Report 74 (EPA 1994). Results from April through June 1984 are included as well (EPA 1986). The locations of air samples with the highest total uranium concentrations were Las Vegas, Nevada; El Paso, Texas; Ross, Ohio; Lynchburg, Virginia; and Phoenix, Arizona (listed in descending concentrations of airborne total uranium). In all cases, atmospheric levels of total uranium were low, in the attocurie/m3 range. The airborne data show 234U to 238U ratios that range from 1.0 to 7.4, many of which are significantly different from the one-to-one ratio found in crustal rock.
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Uranium in airborne dust appears to result from resuspension of soil and, consequently, airborne dust has the same uranium concentration as the soil particles that produce it. Airborne dust near uranium mining or milling operations would be expected to contain higher than background levels of total uranium and have an isotope ratio the same as crustal rock as long as the surface material from which it originated had not experienced significant weathering by moisture. Some examples of airborne uranium levels near mining and milling operations when the industry was actively producing uranium ore are included below for comparison with EPA values in Table 5-2. The annual average concentration of uranium in ambient air taken near the Jackpile Open Pit mine (New Mexico) was 2.4 fCi/m3 (Eadie et al. 1979), and the concentration of uranium in air measured near a Canadian refinery ranged between 1.3 and 134 fCi/m3 (2–200 ng/m3) with a geometric mean of 13 fCi/m3 (20 ng/m3) (Tracy and Meyerhof 1987). Air samples taken near a uranium mill tailings pile showed a uranium concentration of 1 pCi/m3 (NCRP 1984a). Near the Paducah Gaseous Diffusion Plant in Kentucky, where uranium enrichment is performed, the maximum total air alpha activity in 1979 at one location was 0.7 pCi/m3 (UCC 1980).
5.4.2 Water
Until the early 1980s, uranium in drinking water was not often measured except when contamination was suspected. Welford and Baird (1967) found a concentration of 0.02 pCi/L in New York City tap water. UNSCEAR (1977) reported that tap water usually contains less the 0.03 pCi/L.
A large study was performed in which data from the NURE program plus data prepared for the EPA (Drury 1981) were compiled. Over 90,000 water samples were analyzed for uranium. The total data included approximately 35,000 surface water samples that averaged 1.1 pCi/L and approximately 55,000 groundwater samples that averaged 3.2 pCi/L (NCRP 1984a). The population mean was 0.8 pCi/L, which was higher than the 0.03 pCi/L reported by UNSCEAR (1977). Ohanian (1989) reported a population-weighted average concentration of uranium in U.S. community drinking water ranging from 0.3 to 2.0 pCi/L. Another study showed that the average uranium concentrations in drinking water exceeded 2 pCi/L in South Dakota, Nevada, New Mexico, California, Wyoming, Texas, Arizona, and Oklahoma. States in which the average drinking water uranium levels exceeded 1 pCi/L are shown in Figure 5-4 (Cothern and Lappenbusch 1983; EPA 1985j). In another study based on NURE data, the mean uranium concentration in samples of more than 28,000 domestic water supplies was 1.73 pCi/L, with a median concentration range of 0.1–0.2 pCi/L (Cothern and Lappenbusch 1983). The level of uranium
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in 2,228 water supplies was 10 pCi/L or more, while in 979 water supplies, the uranium concentrations were 20 pCi/L or greater. Most of these water supplies were in small towns and served less than a few thousand people (Cothern and Lappenbusch 1983;EPA 1985j).
The EPA ERAMS program measures the uranium content of precipitation. Precipitation samples are only collected during the months of March through May since these spring rain months usually contain the year's highest concentrations of uranium. The data for 234U, 235U, and 238U for 1993 are presented in Table 5-3. The precipitation samples with the highest total uranium concentrations were obtained from Berkeley, California; Niagara Falls, New York; Santa Fe, New Mexico; Wilmington, Delaware; Salt Lake City, Utah; Jacksonville, Florida; and Denver, Colorado (listed in descending concentration of total uranium). In all cases, the uranium concentrations were low, confirming that the atmospheric content of airborne uranium is small.
In some surface waters that have been contaminated by waste discharge and in groundwaters from natural uranium-bearing aquifers, the concentrations of uranium may be higher than the average natural background levels for that area. For example, higher levels of uranium have been observed in water from Ambrosia Lake in New Mexico (uranium milling and mining) (Lapham et al. 1989), the agricultural draining and evaporation pond water of the San Joaquin Valley in California (Bradford et al. 1990), and groundwater from Rocky Flats, Colorado (Laul 1994). The concentration of uranium in creek waters that lead to the Ohio River near the Paducah Gaseous Diffusion Plant in Kentucky ranged from <0.7 to 470 pCi/L (1–700 μg/L) (UCC 1980). Mono Lake, a natural alkaline, saline lake in California, contained 185 pCi/L 238U and 222 pCi/L 234U during the period 1978–80 (Simpson et al. 1982). Analysis of water from the Colorado River and its tributaries during 1985 and 1986 showed that the levels of total uranium ranged from 3.4 to 60 pCi/L (Stewart et al. 1988). In the United States, the highest concentrations of uranium found in surface water and groundwater used as a source of drinking water were 582 and 653 pCi/L, respectively (Drury 1981).
Discharge of dewatering effluents from underground uranium mines and runoff from uranium mine tailings piles have contaminated surface waters and aquifers in New Mexico with elevated levels of gross alpha activity and uranium (NMHED 1989). The concentration of uranium in mine discharge water in New Mexico was 31,500 μg/L (equivalent to 22,680 pCi/L assuming the uranium content is natural uranium)
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(EPA 1985j). Groundwater from an aquifer adjacent to a uranium mill tailings pile in Falls City, Texas, was also found to have concentrations of uranium above natural background levels (DOE 1994).
The concentrations of 234U and 238U in groundwater from Cambrian-Ordovician sandstone aquifers in Illinois range from 40. The lowest ratios were found in unconfined aquifers in primary recharge zones while ratios >20 were found in the confined zones aquifer. It was suggested that glacial recharge in unconfined zones might be responsible for the high 234U to 238U ratios (Gilkeson and Cowart 1987). Fifty-five groundwater samples from the Lockatong and Passaic Formation in the Newark Basin in New Jersey were analyzed during 1985–1987. These samples were found to contain 0.1–40 pCi/L total uranium, with a median value of 2.1 pCi/L (Szabo and Zepecza 1987). Uranium concentrations measured in 7 samples of groundwater from the Raymond Basin in California ranged from 5.3 to 43.7 pCi/L (Wiegand et al. 1987).
Water in a private well in Maine, thought to be of geologic origin, was reported to contain as much as 403 μg/L uranium (approximately 270 pCi/L) (Lowry et al. 1987). Elevated levels of uranium measured in waters from private wells in northern and northeastern Nebraska were thought to be due to the upward migration of uranium from bedrock and heavy use of phosphate fertilizers. Uranium values up to 110 pCi/L were measured (NEDH 1989). The concentrations of uranium in U.S. groundwaters were estimated using a conceptual model based on the geochemical and hydrological characteristics of aquifers.
The population-weighted average uranium concentration in groundwaters used as sources of drinking water in all 50 states was found to range from 0.05 to 4.6 pCi/L, with a mean value of 0.55 pCi/L (Longtin 1988). This mean is lower than the population-weighted uranium value for finished waters of 0.8 pCi/L (NCRP 1984a). Some methods which may be suitable for reducing the concentration of uranium in drinking water include lime softening, coagulation/precipitation, and filtering; however, these methods may not efficiently remove the uranium.
5.4.3 Soil
Table 5-4 shows the average concentrations of uranium in several types of rocks and soils (NCRP 1984a). The radioactivity in soils is similar to that in the rocks, usually bedrock, from which it derives. The average soil concentration of 234U from Table 5-4 is 0.6 pCi/g. Since the activity of 234U accounts for
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approximately one-half of the total activity in natural uranium (see Chapter 3), the value in Table 5-4 may be multiplied by two to obtain the total uranium in soils (approximately 1.2 pCi/g).
There are wide variations from the values presented in the table, particularly in areas where uranium minerals are more concentrated. Concentrations of uranium in Louisiana soils ranged from 2.35 to
3.98 μg/g (1.6–2.7 pCi/g) (Meriwether et al. 1988), while uranium concentrations in phosphate rock in north and central Florida ranged from 4.5 to 83.4 pCi/g (6.8–124 μg/g) (EPA 1985j).
Soil samples adjacent to Los Alamos, New Mexico, taken during 1974–1977 contained total uranium in the range of 0.1–5.1 μg/g (0.067–3.4 pCi/g), with a mean concentration of 1.6 pCi/g (2.4 μg/g) (Purtymun et al. 1987). The concentrations of uranium in soils adjacent to the Hanford Fuel Fabrication Facility near Richland, Washington, that were collected during 1978–81 ranged from 0.51 to 3.1 pCi/g (0.76–4.6 μg/g), with a median value of 1.2 pCi/g (1.8 μg/g). The control samples for the Hanford Fuel Fabrication Study contained uranium at concentrations of 0.21–0.86 pCi/g (0.32–1.128 μg/g), with a median value of
0.49 pCi/g (0.73 μg/g) (Price and Kinnison 1982). Uranium in the soil within the property boundary of the Paducah Gaseous Diffusion Plant in Kentucky ranged from 3.3 to 4.8 pCi/g (4.9–7.1 μg/g), whereas off-site samples taken as far as 12 miles away contained uranium at levels of 3.8–6.0 pCi/g (6.4–9.0 μg/g) (UCC 1980). Soil monitoring data from the area surrounding the Feed Material Production Center at Fernald, Ohio, showed that the uranium concentrations within an 8-km2 area were between 3 pCi/g and 23 pCi/g (4.5–34 μg/g) compared to an mean of 2.2 pCi/g (3.3 μg/g) for natural background levels (Stevenson and Hardy 1993). Other investigators have detected uranium levels in surface soils at the Fernald site as high as 50 times natural background levels (Miller et al. 1994).
5.4.4 Other Environmental Media
Concentrations of uranium have been determined in meat and fish (Table 5-5). The uranium content measured in tissues of cattle herds grazing in pastures close to the Rocky Flats Plant in Colorado was slightly higher than in other cattle, reflecting possible contamination from this source (Smith and Black 1975). The average concentrations of uranium in game fish (surface feeders) collected from a reservoir at locations upstream and downstream from the Los Alamos National Laboratory were 2.9 ng/g dry weight (dw) (0.0019 pCi/g) and 4.9 ng/g dw (0.0033 pCi/g), respectively (Fresquez et al. 1994). The corresponding values in nongame, bottom-feeding fish were 7.9 ng/g dw (0.0058 pCi/g) and 17.7 ng/g dw
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(0.012 pCi/g), respectively. The concentrations of uranium in fish muscle (dw) from a Canadian lake receiving uranium mill effluents were 7–11 times higher than in fish caught in uncontaminated lakes, but this uranium may have only been attached to the gills (Swanson 1985).
The mean uranium concentration in vegetation from Ambrosia Lake, New Mexico, (site of mining and milling activities) was measured at 0.3 pCi/g dw compared to 4 fCi/g dw for vegetation from a control site (Lapham et al. 1989). Although the concentrations of uranium in muscle from exposed cattle were indistinguishable from uranium levels in muscle from control cattle, levels of uranium in liver and kidney tissues were 4 times higher in exposed cattle than in control cattle, and levels of uranium in femur samples were 13 times higher than in controls, indicating that kidney and liver slightly bioconcentrate uranium while muscle does not (Lapham et al. 1989).
5.5 GENERAL POPULATION AND OCCUPATIONAL EXPOSURE
General population exposure to uranium occurs by ingestion of food and drinking water and by inhalation of air. The pathways are shown in Figure 5-3.
Table 5-5 depicts uranium levels in various types of food in the United States. Measurements of normal levels of dietary 234U and 238U indicate that foods consumed contain about 0.3–0.5 pCi/day for each uranium isotope (0.6–1.0 pCi/day [0.9–1.5 μg/day] total uranium) (EPA 1985j; Welford and Baird 1967). Based on consumption rates, root crops such as potatoes, parsnips, turnips, and sweet potatoes contribute approximately 38% of total dietary intake of uranium (EPA 1985j).
Ingestion of food grown in the vicinity of a uranium mill may lead to an intake up to 3 pCi/day uranium (Rayno 1983). Other investigators have estimated a dietary intake of 2.86–4.55 mg/day for individuals living near a uranium mine (Yamamoto et al. 1971).
An alternate method for estimating uranium intake is to measure the daily excretion of uranium in urine and feces. Using this method in a study of 12 subjects in Utah, it was estimated that the average dietary intake for the Salt Lake City population was 4.4±0.6 μg, an intake that is higher than that reported for New York City, Chicago, and San Francisco residents (1.3–1.4 μg) (Singh et al. 1990).
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Intakes of uranium in food may also increase when certain ceramic glazed dishes are used for serving or storing food (Landa and Councell 1992). Leaching occurs on contact with acidic foods or beverages. Experiments show that when a ceramic glazed plate was kept in contact with a 4% acetic acid solution for 24 hours, the concentration of uranium in the leachate was 31.8 mg/L (Landa and Councell 1992).
Uranium glazed commercial ceramic dinnerware is no longer made and sold because it was determined that the uranium is leachable by acidic foods and beverages (Landa and Councell 1992). Experiments show that when a Fiesta tableware plate was kept in contact with 20 mL of 4% acetic acid solution for 24 hours, the quantity of uranium in the leachate was 600 μg (400 pCi). Other liquids were much less effective at leaching uranium, with water giving a value over three orders of magnitude lower, and other uranium glazed ceramics were much less leachable (Landa and Councell 1992).
Uranium intakes from food in Japanese diets from two control areas ranged from 0.86 to 1.02 μg/day (Yamamoto et al. 1971). A more recent study reported a mean value of 0.71 μg/day for Japanese males from 31 prefectures (Shiraishi et al. 1992). Worldwide intake values for uranium have been reported at an average of 1 pCi/day (1.5 μg/day) (range 0.6–3.2 pCi/day [0.9–4.8 μg/day]) (Linsalata 1994).
Concentrations of uranium from selected drinking water supplies in the United States were analyzed by the EPA laboratories and found to be generally <1 pCi/L (EPA 1985j). Based on data obtained from the NURE program plus data prepared for the EPA (Drury 1981), a population-weighted average of 0.8 pCi/L uranium was determined. In another study, Ohanian (1989) reported population-weighted average concentrations of uranium in U.S. community drinking water ranging from 0.3 to 2.0 pCi/L. Considering an individual water intake of approximately 1.7 L/day, and an average intake of uranium from drinking water of 0.8 pCi/L as reported in the EPA study, the total intake of uranium for an individual from drinking water each day is approximately 1.4 pCi.
Uranium is also taken into the body by the inhalation route. The average daily intake of uranium from inhalation of air has been estimated to range from 0.007 pCi/day (0.010 μg/day) (Cothern 1987) to 0.0007 pCi/day (0.0010 μg/day) (UNSCEAR 1988). This value may be somewhat higher for persons living near sources of uranium emission. Glass makers and potters who use uranium-containing enamels may be exposed to small amounts of uranium from handling the powder or from fuming operations in glass making (Rossol 1997). In general, however, exposure to uranium from inhalation is small compared to exposure from food and drinking water.
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Measurements of concentrations of uranium have been made in human tissues and body fluids resulting from consumption of food and water and from natural background sources. These are non-occupationally exposed populations. Two longtime residents of Los Alamos, New Mexico (one a smoker and one not) were shown to have uranium tissue concentrations for the skeleton (average 5.8 μg/g wet weight) and liver (average 0.08 μg/kg) in closer agreement with the Reference Man (Kathren 1997; ICRP 1975) than those reported in New York City residents (Fisenne and Welford 1986). Values of uranium in whole blood measured in New York City residents and Illinois residents averaged 0.14 μg/kg (0.09 pCi/kg) and
0.1 μg/kg (0.07 pCi/kg), respectively, compared to a mean value worldwide of 0.58 μg/kg (Fisenne 1988). Mean concentrations of uranium were measured in the organs of persons representing all age groups fromdifferent parts of the United States. The uranium values for lungs, liver, kidney, and bone (vertebrae, rib,and skeleton) were 0.5–1.17 μg/kg (0.34–0.78 pCi/kg), 0.12–0.33 μg/kg (0.08–0.22 pCi/kg),0.39–1.00 μg/kg (0.26–0.67 pCi/kg), and 0.25–1.9 μg/kg (0.17–1.3 pCi/kg), respectively (Fisenne et al.1988; Fisenne and Welford 1986; Singh et al. 1986b). These differences reflect dietary variations.
Workers engaged in the extraction and processing of uranium are occupationally exposed to uranium. Industries where uranium exposures are known to have occurred are uranium mining and milling, uranium conversion and enrichment, uranium fuel fabrication, and nuclear weapons production. Epidemiologic surveys were initiated in the United States as early as 1950 to study the effects of uranium exposure on uranium millers, and similar studies were performed of workers at the Oak Ridge Gaseous Diffusion Plant in Oak Ridge, Tennessee, where uranium conversion and enrichment were performed. Those studies attributed the health decrement to radon progeny and other toxicants and not directly to the uranium (BEIR IV 1988).
5.6 EXPOSURES OF CHILDREN
This section focuses on exposures from conception to maturity at 18 years in humans and briefly considers potential pre-conception exposure to germ cells. Differences from adults in susceptibility to hazardous substances are discussed in Section 2.6, Children’s Susceptibility.
Children are not small adults. A child’s exposure may differ from an adult’s exposure in many ways. Children drink more fluids, eat more food, and breathe more air per kilogram of body weight, and have a larger skin surface in proportion to their body volume. A child’s diet often differs from that of adults. The
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developing human’s source of nutrition changes with age: from placental nourishment to breast milk or formula to the diet of older children who eat more of certain types of foods than adults. A child’s behavior and lifestyle also influence exposure. Children crawl on the floor; they put things in their mouths; they may ingest inappropriate things such as dirt or paint chips; they spend more time outdoors. Children also are closer to the ground, and they do not have the judgement of adults in avoiding hazards (NRC 1993).
Specific information on the exposure of children to uranium is not available. As for adults in the general population, small exposures occur from normal ingestion of food and drinking water and inhaling air. These exposures may be higher in areas with naturally high uranium soil levels or near uranium processing sites and hazardous waste sites containing uranium. A study of uranium content in bone from three age groups (99%) in the form of insoluble oxides of uranium which have very low bioavailability.
As for adults, the potential for uranium exposure is greater for children who consume foods grown in areas with elevated concentrations of uranium in the soil and for children with elevated concentrations of uranium in their drinking water (EPA 1985; NCRP 1984a). Other home exposures are unlikely since no household products or products used in crafts, hobbies or cottage industries contain significant amounts of uranium, except in cases where uranium-bearing rocks are used in and around the home for decorative, collection, or construction purposes (ATSDR 1997)
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No information is available on whether children differ from adults in their weight-adjusted intake of uranium. The fractional absorption of uranium (as uranyl nitrate and uranyl citrate) by the oral route was higher in neonatal than in adult rats and swine (Sullivan 1980b; Sullivan and Gorham 1982). In a mathematical model developed by the International Commission for Radiological Protection (ICRP) for risk assessment, one of the assumptions is that the fractional absorption of ingested uranium is twice as high in children under the age of 1 year compared to adults.
Uranium exposure to children from parents’ work clothes, skin, hair, tools, or other objects from the workplace is possible if the parent uses uranium at work. However, in a comprehensive review of incidents of home contamination by workers (NIOSH 1997), no cases of uranium contamination were described.
As a radionuclide, uranium is potentially genotoxic and thus it is important to know if parental exposure to uranium could affect the developing fetus or germ cells. However, epidemiological studies of workers exposed to uranium show no evidence of genotoxic effects. This is most likely due to the very low specific activity, the low systemic absorption of uranium, and the lack of concentration of uranium in the germ cells. Genotoxic effects to parental germ cells or to a developing fetus are not likely at probable levels of exposure to uranium from the environment or at hazardous waste sites. Some uranium is stored in bone, but it is not known if this uranium is released during pregnancy and lactation, when it could result in exposure to the fetus or infant.
5.7 POPULATIONS WITH POTENTIALLY HIGH EXPOSURES
Higher rates of uranium exposure have been reported for some populations. The potential for uranium exposure is greater for individuals who consume foods grown in areas with elevated concentrations of uranium in the soil, and for individuals with elevated concentrations of uranium in their drinking water (EPA 1985; NCRP 1984a). Industries where higher exposures to uranium are known to occur include uranium mining and milling, uranium conversion and enrichment, uranium fuel fabrication, and nuclear weapons production (BEIR IV 1988; Miller 1977; NCRP 1984a; West et al. 1979). Other groups with the potentially higher exposures include persons involved in producing and using phosphate fertilizers and individuals living and working near fossil fuel plants (Jaworowski and Grzybowska 1977; NCRP 1984a; Tadmor 1986; Weissman et al. 1983). Uranium compounds were previously used in dental appliances, and individuals with dental work of this kind have potentially higher exposures (Sairenji et al. 1980).
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5.8 ADEQUACY OF THE DATABASE
Section 104(i)(5) of CERCLA, as amended, directs the Administrator of ATSDR (in consultation with the Administrator of EPA and agencies and programs of the Public Health Service) to assess whether adequate information on the health effects of uranium is available. Where adequate information is not available, ATSDR, in conjunction with NTP, is required to assure the initiation of a program of research designed to determine the health effects (and techniques for developing methods to determine such health effects) of uranium.
The following categories of possible data needs have been identified by a joint team of scientists from ATSDR, NTP and EPA. They are defined as substance-specific informational needs that if met would reduce the uncertainties of human health assessment. This definition should not be interpreted to mean that all data needs discussed in this section must be filled. In the future, the identified data needs will be evaluated and prioritized, and a substance-specific research agenda will be prepared.
5.8.1 Identification of Data Needs
Physical and Chemical Properties. Pertinent data on the physical and chemical properties of uranium and uranium compounds are available in the literature.
Production, Import/Export, Use, Release, and Disposal. Data regarding the past and present production (ABMS 1994; EPA 1985a) and import/export volumes (USDOC 1995) for uranium are available. The uses of uranium and uranium compounds are well known (EPA 1985j; Stokinger 1981; USDI 1980). Other than glazed ceramic foodware and decorative items (Landa and Councell 1992) and dental appliances (Sairenji et al. 1980), consumer contact with uranium products is negligible. Since uranium is not covered under SARA, Title III, manufacturers and users are not required to report releases to the EPA. There is a lack of data on the release and disposal of uranium during mining, milling, and chemical processing and its use during fuel cycle operations. The disposal of uranium is governed by the
U.S. Nuclear Regulatory Commission (NRC) regulations (10 CFR 61), and releases of uranium to the environment are governed by NRC and EPA regulations (10 CFR 20, Appendix B; 40 CFR 190; 40 CFR 192). Since significant amounts of depleted uranium are used on modern battlefields, it would be useful to have more information on the export of depleted uranium to other nations, the disposal of related wastes in
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the United States, and the mass of depleted uranium released to long-distance air transport when projectiles are used against different target types.
Environmental Fate. For solids, there is a need to determine uptake factors into edible portions of plants and not just adherance to the root structure. For the solid-liquid interface, a method is needed to determine method by which 234U to 238U ratio deviates from unity such that the EPA ERAMS water sample results indicate disequilibrium. Uranium enters the atmosphere in particulate form from natural sources and from uranium mining, milling, and processing. Dry or wet deposition from the atmosphere to soil and water can occur (Essien et al. 1985). Little experimental data on the particle size and residence time of uranium and uranium compounds present in ambient atmospheres are available. Additional data regarding the measured particle size of uranium compounds in ambient air, settling velocity, and knowledge of the chemical forms and lifetime of the particles in air would be useful. Although recent studies have characterized the oxidation states and chemical forms of some uranium compounds (UO2 and UO3) (Dodge and Francis 1994; Wersin et al. 1994), more data identifying the chemical forms of uranium in the environment are needed to better understand the fate and transport of uranium. Since significant amounts of depleted uranium are used on modern battlefields, it would be useful to have more information on the export of depleted uranium to other nations and the disposal of related wastes in the United States, as well as estimates of projectile quantities that aerosolize to a significant extent and associated downwind air contamination levels.
Bioavailability from Environmental Media. The absorption and distribution of uranium as a result of inhalation and ingestion exposures have been discussed in Sections 2.3.1 and 2.3.2. However, quantitative data relating to physical/chemical properties such as particle size, chemical form, and degree of absorption with the bioavailability of uranium from inhaled air particles and inhaled and/or ingested soil particles, are lacking. Such studies would be useful in assessing potential public health impact of uranium to people living near a hazardous waste site.
Food Chain Bioaccumulation. Information about uranium bioaccumulation in fish (Mahon 1982; Poston 1982; Swanson 1983; Waite et al. 1988) is available. Also data concerning levels of uranium in various foods (EPA 1985j) are available. These data indicate that uranium does not biomagnify in the food chain (Ahsanullah and Williams 1989; Morishima et al. 1977; Swanson 1983, 1985). Data on the levels of
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uranium in food grown in contaminated areas are limited. Additional data are needed on whether the uptake of uranium in fish is restricted to the gills and how much actually distributes to the meat.
Exposure Levels in Environmental Media. Reliable monitoring data for the levels of uranium in contaminated media at hazardous waste sites are needed so that the information obtained on levels of uranium in the environment can be used in combination of the known body burden with uranium to assess the potential risk of adverse health effects in populations living in the vicinity of hazardous waste sites.
The levels of uranium in airborne particles and precipitation have been monitored since 1973 (EPA 1986, 1994). Data from several large studies of uranium in domestic water supplies are available (Cothern and Lappenbusch 1983; Drury 1981), as are data from studies of groundwater and surface water (NCRP 1984a). The primary source of recent information on the occurrence of uranium in drinking water is the National Inorganics and Radionuclides Survey (NIRS) conducted by EPA (EPA 1991). Some monitoring data are available for uranium-contaminated soils and sediments associated with the nuclear fuel cycle. Better information on background levels in the environment and speciation of uranium in soils and sediments would be useful for determining which species lead to actual public exposure.
Exposure Levels in Humans. Although some data on the levels of uranium in human tissues exposed to natural background levels (food, water, and air) are available, few measurements have been made on the uranium content in human tissues. The principal source of information about occupationally exposed individuals is the U.S. Transuranium and Uranium Registries (USTUR) Tissue Program and database, established to document uranium levels and distribution in human tissues for occupationally exposed workers (PNL 1981). Several major database files are available. The Radiochemical file contains information about radiochemical analysis of tissue donations from occupationally exposed individuals. The Health Physics file contains bioassay and other health physics data. These two databases are regularly updated. The Medical file contains abstracted personal, medical, and clinical data; the Pathology file contains autopsy and pathology information; and the Skeletal Estimate file contains estimated actinide concentrations for unanalyzed half skeletons from donors (USTUR 1999).
Exposures of Children. Children will be exposed to uranium in the same manner as adults in the general population (i.e. ingestion of food and water and inhalation of air). A study of uranium content in bone from three age groups (<13, 13–20, 20–25 years old) reported somewhat higher uranium content in the
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youngest compared to the oldest age group (approximately 1.5–3 fold); however, there were only 2–4 subjects in each group and the results were not statistically significant (Broadway and Strong 1983). Since the skeletons of children are growing (higher rate of bone formation), it is possible that a higher fraction of circulating uranium will be deposited in bone than in adults. Further information is needed on bone levels of uranium in children to determine if this is the case. Uranium is found in all soil, and at potentially higher levels at some hazardous waste sites. Since children may have oral exposure to soil through hand-to-mouth activity, bioavailability studies of uranium in soil may be useful to assess the risk of this type of exposure. There is some evidence that neonatal animals absorb uranium in the gastrointestinal tract to a greater extent than adults. Experiments to confirm this finding and to determine how long into maturation a difference exists would help refine risk assessment for uranium exposure in children.
Exposure Registries. A voluntary exposure registry, the U.S. Transuranium and Uranium Registry (USTUR) for occupationally exposed individuals, was established at Richland, Washington, in 1968 as the National Plutonium Registry for investigation of the potential hazards for occupational exposure to uranium. In 1971, additional radiochemistry support was provided by Los Alamos National Laboratory. The United States Uranium Registry was created as a separate entity in 1978, and the two registries operated in parallel until 1987, when a single director was given responsibility for both registries. In 1992, the management and operation of the registries was combined at Washington State University under a grant from the U.S. Department of Energy. The primary goals are to develop information on the distribution and dose of uranium and transuranic elements in humans, providing data for verification or development of radiation protection standards, and to determine and evaluate health effects due to exposure to these radioactive elements.
5.8.2 Ongoing Studies
The Federal Research in Progress database lists ongoing studies about environmental effects of uranium (FEDRIP 1999). These are shown in Table 5-6.
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The purpose of this chapter is to describe the analytical methods that are available for detecting and/or measuring and monitoring uranium in environmental media and in biological samples. The intent is not to provide an exhaustive list of analytical methods that could be used to detect and quantify uranium. Rather, the intention is to identify well-established methods that are used as the standard methods of analysis. Many of the analytical methods used to detect uranium in environmental samples are the methods approved by federal agencies such as EPA, DOE, and the National Institute for Occupational Safety and Health (NIOSH). Other methods presented in this chapter are those that are approved by a trade association such as the Association of Official Analytical Chemists (AOAC) and the American Public Health Association (APHA). Additionally, analytical methods are included that refine previously used methods to lower detection limits, and/or to improve accuracy and precision.
Most of the equipment and analytical methods described in this chapter for field measurements and, to a lesser extent, laboratory sample analysis are summarized in the Multi-Agency Radiation Survey and Site Investigation Manual (MARSSIM 1997). It is anticipated that its companion manual, the Draft Multi-Agency Radiological Laboratory Analytical Protocols (MARLAP) manual, will robustly describe relevant analytical equipment and methods, and be available for public comment in 2000.
6.1 BIOLOGICAL MATERIALS
Uranium can enter the human body through inhalation, ingestion, or penetration through the skin. Measurement of the quantities of uranium in the body can be performed by two primary methods, in vivo measurements and in vitro measurements. These types of measurements are called bioassays. In vivo techniques measure the quantities of internally deposited uranium directly using a whole body counter while in vitro techniques permit estimation of internally deposited uranium by analysis of body fluids, excreta, or (in rare instances) tissues obtained through biopsy or postmortem tissue sectioning (NCRP 1987) (USUTR 1999). Some of these analytical methods are summarized in Table 6-1.
6.1.1 Internal Uranium Measurements
In vivo or direct measurements of uranium in the body are made with radiation detector systems and associated electronics called whole body counters that measure radiation as it leaves the body from internally deposited uranium. In vivo assays are the most direct method of quantifying internally deposited
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radioactive materials. However, not all radionuclides emit radiations than may be detected outside the body (234U and 238U, for example) (NCRP 1978). The most commonly used detectors for uranium in vivo counting are sodium iodide, phoswich (NaI and CsI sandwich), and hyperpure germanium which measure the gamma rays emitted during uranium decay (DOE 1988). Since the gamma radiations emitted from uranium and a number of its progeny are the same as those emitted by uranium in the environment, shielded rooms are normally used to house the uranium internal monitoring equipment to ensure that background radiation is as low as possible (DOE 1999; Parrington et al. 1996) . Although whole body counters may be made in many configurations, a chest counter is usually used for inhaled uranium. In vivo analysis is widely used throughout the nuclear industry, both commercial and government, for quantifying levels of insoluble uranium in the body. In vitro analysis (see Section 6.1.2) is often used in conjunction with whole body counting for monitoring workers handling uranium (DOE 1988).
In vivo counting systems are calibrated using tissue-equivalent phantoms. These phantoms have shapes similar to the human torso and are made of polystyrene or other tissue equivalent material. Standard uranium sources of known activity are inserted into the phantom at locations where uranium would be expected to accumulate in a human body (DOE 1988). Relationships are determined between the uranium activity measured by the detection system and the known activity in the phantom (DOE 1988; HPS 1996).
There are limitations associated with in vivo counting uranium measurements. First, soluble uranium is readily excreted, with fractions retained for varying periods in the bone and kidney, so detectability depends on factors such as intake quantity, chemical and physical form, biodistribution fraction, time since intake, background uranium contribution, analysis time, and detection system efficiency. Second, only the 235U isotope can be detected using the sodium iodide or hyperpure germanium detectors, since 234U and 238U decay does not result in emission of gamma rays, which are required for detection by sodium iodide and hyperpure germanium detectors (NCRP 1987). In such cases, indirect in vitro methods can be used for measuring uranium in urine or feces (DOE 1988; HPS 1996). Analytical equipment and procedures vary widely among laboratories and often require individual-specific input (NCRP 1987). The Minimum Testing Level (MTL) of 0.81 nCi 235U (lung) has been established as a performance level to which laboratories are expected to adhere for in vivo detection (HPS 1996).
6.1.2 In Vivo and In Vitro Uranium Measurements
In vitro uranium analyses are routinely performed in support of a personnel monitoring program, or in cases where the size of an operation does not justify the cost of whole body counter facilities. These analyses are usually done on urine samples, but other types of body materials may also be used (e.g., feces or blood).
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Urinalysis is effective for analysis of transportable or soluble uranium. A fraction of insoluble uranium also appears in the urine (DOE 1988).
The excretion of uranium in fecal material results primarily from intakes by ingestion, and includes uranium swallowed after inhalation. Usually, uranium will appear in feces within hours after intake thus providing a rapid means of determining whether an intake has occurred. Fecal analysis requires prechemistry preparation that includes ashing of the sample, cleaning by co-precipitation, and solvent extraction followed by electrodeposition. Alpha spectroscopy is then performed (Singh and Wrenn 1988). Urinalysis is typically favored over both fecal and blood analysis because it is generally more sensitive and less costly, and because fecal analysis provides no uptake or retention information and blood analyses is invasive.
Several methods that do not require chemical separation are available for measuring uranium in urine (in units of total mass or total activity). These methods include spectrophotometric (total mass), fluorometric (total mass), kinetic phosphorescence analysis (KPA) (total mass), and gross alpha (total activity) analyses (Wessman 1984). The most widely used methods for routine uranium analysis are α-spectrometry and liquid scintillation spectrometry. These methods utilize the natural radioactivity of uranium and are sensitive and require little sample preparation. Photometric techniques such as fluorometry and phosphorometry are less widely used, but kinetic phosphorescence analysis is becoming more widely used. Measurements of total uranium do not provide the relative isotopic abundance of the uranium isotopes, but this may only be important when converting between activity and mass when the isotopic ratios are uncertain.
If quantification of an individual uranium isotope is needed (e.g., 234U, 235U, or 238U), the most commonly used methods require chemical separation followed by α-spectrometry, or chemical separation and electrodeposition followed by α-spectrometry (see Table 6-1). Mass spectrometric methods have emerged as sensitive, reliable techniques for determining uranium isotopes at low concentrations. Inductively coupled plasma-mass spectrometry (ICP-MS) requires sample preparation, but is rapid and is becoming less expensive (Twiss et al. 1994).
Uranium may also be measured in fecal material using the same methods identified above for urinalyses, except that this matrix requires extensive preparation. For α-spectroscopy, this includes ashing of the sample, cleaning by co-precipitation, and solvent extraction followed by electrodeposition and α-spectroscopy (Singh and Wren 1988). In the other methods, electrodeposition is replaced with an equipment-specific step, such as direct injection for ICP-MS and mixing with a scintillation cocktail for liquid scintillation. Table 6-1
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The MTL for 234U, 235U, and 238U using α-spectroscopy is 0.54 pCi/L in urine. An acceptable minimum detection activity of 20 μg/L of urine has also been established for natural uranium based on mass determination (HPS Standard N13.30 1996). Determining the accuracy and precision of the quantification methods for biological materials by either in vivo or in vitro methodologies requires that standard, certified sources with known concentrations of appropriate radionuclides be available for calibrations. The primary source of certified standards is the National Institute of Standards and Technology (NIST) (Inn 1987). An aqueous solution of uranium containing 10 mg/mL (SRM 3164) standard stock solution is available, as are solutions of 232U (1.1 nCi/g [40 Bq/g]) (SRM 4324) and 238U, "natural uranium," (6.7 nCi/g [250 Bq/g]) (SRM 4321B) (NIST 1995). Standard Reference Materials of human lung (SRM 4351) and human liver (SRM 4352) are also available from NIST.
6.2 ENVIRONMENTAL SAMPLES
Two types of methods are commonly used for measurement of uranium in environmental samples. The first are field surveys using portable survey instruments, and the second is analysis of samples procured in the field that are returned to the laboratory for quantification.
6.2.1 Field Measurements of Uranium
Uranium measurements in the field are typically qualitative in nature in that the instruments simply respond to alpha emissions, regardless of their isotopic origin. However, the levels can be measured quantitatively if key parameters are known, such as relative abundances of all alpha-emitting isotopes present, the thickness of the layer being assessed, and the detection efficiency of the instrument for the type of surface being assessed. Measurements in the past have typically been made using a portable, hand-held alpha scintillation detector (e.g. ZnS) equipped with a count rate meter, which detects alpha radiation while discriminating against beta-emitters in the same area. However, the need for low detection limits in radiological remediation efforts has found a more suitable and sensitive instrument in the large-area gas-flow proportional counter. These instruments can be carried by an individual or attached to a holder for maintaining a selected surface-to-detector distance. The latter method can be integrated into a system which moves along a surface at a predetermined velocity recording spatially-related real-time data for later graphical imaging of absolute surface activity distributions (DOE 1988). These surveys can also be performed on people whose skin or clothing is contaminated. Survey instruments can provide a quick
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estimate or a measure of the level of activity that might be present. However, more accurate measurement of uranium activity may require that samples be taken for laboratory analyses. Under normal usage, the lowest level of uranium that can be reliably detected using an alpha scintillation survey meter is 200–500 disintegrations per minute/100 cm2 (0.09–0.23 nCi/100 cm2) (DOE 1988); however, detection of levels several time lower is practical with gas flow proportional counters, especially when used in the integrate mode. Detection capability varies with the type of detector used, the active area of the probe, the electronics, etc.
Several limitations are associated with the measurement of uranium by portable survey instruments. First, the uranium must be present on the surface of the material being surveyed. Since uranium decays by emission of α particles, which travel only short distances in materials, any uranium that is imbedded in the surface being surveyed will be partially or completely masked . Secondly, when performing surveys, it must be possible to place the detector very close to the surface being surveyed (i.e., approximately one-quarter of an inch) (DOE 1988, 1994), and uneven surfaces that are unintentionally touched can tear the detector window, disabling the instrument. Additional information is available in MARSSIM (1997) on the use and usefulness of field survey instruments.
6.2.2 Laboratory Analysis of Environmental Samples
Analytical methods for measuring uranium in environmental samples are summarized in Table 6-2. The available methods can be divided into two groups: chemical methods to determine the total mass of uranium in a sample, and radiological methods to determine amounts of individual isotopes. Environmental media that have been tested for uranium include air filters, swipes, biota, water, soil, and others; a full range of laboratory analysis methods has been used to quantify the total uranium or its individual isotopes. The equipment and methods tend to improve over time. The radiological analysis methods primarily use high resolution α-spectroscopy, although gamma spectroscopy is usable with great care. The chemical methods which are often used include spectrophotometry, fluorometry, and kinetic phosphorescence, with the recent addition of various mass spectrometer applications (ICP-MS, AES-MS, and accelerator-MS). If conversions between mass and activity are to be made accurately, prior knowledge of the relative abundance of the various uranium isotopes must be available or measured radiologically. A few media-specific methods which have been used successfully for measuring uranium concentrations in environmental samples are described below. The current trend, however, is away from prescriptive methods and toward
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performance-based methods which enable the user to optimize their available analytical tools. A cornerstone of this method is the development of Data Quality Objectives and the use of Data Quality Assessment to ensure that the selected method is properly developed and the results are of the appropriate quality (DOE 1997; EPA 1994b, 1996).
DOE’s method for analyzing environmental materials is based on a method of Welford et al. (1960) and involves preparing triplicate air, water, and soils samples by concentrating or isolating uranium from the media prior to separation in an anion exchange column, followed by fluorometric analysis (DOE 1997).
In one analytical method for air filters, the air filters are ashed, silica content is volatilized with hydrogen fluoride, uranium is extracted with triisooctylamine, purified by anion exchange chromatography and coprecipitated with lanthanum as fluoride. The precipitated uranium is collected by filtration, dried, and α-spectroscopy is performed (EPA 1984b). The activities of 234U, 235U, and 238U are determined based on the number of counts that appear in the α energy region unique to each isotope. This method is used by the EPA National Air and Radiation Environmental Laboratory for measurement of uranium in air as part of the Environmental Radiation Ambient Monitoring System (see Chapter 5).
Singh and Wrenn (1988) describe a method for uranium isotopic analysis of air filters. Air filters are ashed, redissolved, and co-precipitated with iron hydroxide and calcium oxalate. The uranium is further purified by solvent extraction and electrodeposition. An alpha spectroscopy detection level of 0.02 dpm/L for 238U in solution was reported (Singh and Wrenn 1988).
Considerable work has been done to develop methods for analysis of uranium in water. In 1980, the EPA published standardized procedures for measurement of radioactivity in drinking water which included uranium analysis by both radiochemical and fluorometric methods (Krieger and Whittaker 1980), and more recently developed an ICP-MS method. An example of each is provided below.
The radiochemical method quantifies gross α activity utilizing either a gas flow proportional counter or a scintillation detection system following chemical separation. In the EPA radiochemical method, the uranium is co-precipitated with ferric hydroxide, purified through anion exchange chromatography, and converted to a nitrate salt. The residue is transferred to a stainless steel planchet, dried, flamed, and counted for α particle activity (Krieger and Whittaker 1980).
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For the fluorometric method, uranium is concentrated by co-precipitation with aluminum phosphate, dissolved in diluted nitric acid containing magnesium nitrate as a salting agent, and the co-precipitated uranium is extracted into ethyl acetate and dried. The uranium is dissolved in nitric acid, sodium fluoride flux is added, and the samples fused over a heat source (EPA 1980).
The ICP-MS method was developed for measuring total uranium in water and wastes. The sample preparation is minimal—filtration for dissolved uranium, acid digestion for total recoverable uranium. Recovery is quantitative (near 100%) for a variety of aqueous and solid matrices and detection limits are low, 0.1 μg/L for aqueous samples and 0.05 mg/kg for solid samples (Long and Martin 1991).
The EPA developed two methods for the radiochemical analysis of uranium in soils, vegetation, ores, and biota, using the equipment described above. The first is a fusion method in which the sample is ashed, the silica volatilized, the sample fused with potassium fluoride and pyrosulphate, a 236U tracer is added, and the uranium extracted with triisooctylamine, purified on an anion exchange column, coprecipitated with lanthanum, filtered, and prepared in a planchet. Individual uranium isotopes are separately quantified by high resolution alpha spectroscopy and the sample concentration calculated using the 236U yield. The second is a nonfusion method in which the sample is ashed, the silica volatilized, a 236U tracer added, and the uranium extracted with triisooctylamine, stripped with nitric acid, co-precipitated with lanthanum, transferred to a planchet, and analyzed in the same way by high resolution α-spectroscopy (EPA 1984).
The detection capability of any measurement process is an important performance characteristics, along with precision and accuracy. The Lower Limit of Detection (LLD) has been adopted to refer to the intrinsic detection capability of the measurement process (sampling through data reduction and reporting) (USNRC 1984). Factors that influence the LLD include background count rate, sensitivity of detector, and, particularly, the length of time a sample and background are counted. Because of these variables, LLDs between laboratories, employing the same or similar chemical separation procedures, will vary. Additional examples of the techniques for quantification of uranium (as described above) are available, as well as examples of less frequently used techniques. These are identified in Table 6-3.
Determining the accuracy of the analytical methods for environmental samples and for calibrating radiation instrumentation requires that standard, certified radioactive sources with known concentrations of uranium,
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or other appropriate radionuclides, be available for use. The primary source of such certified standards is NIST (Inn 1987). An aqueous solution of uranium containing 10 mg/mL (SRM 3164) standard stock solution is available, as are solutions of 232U (1.1 nCi/g [40 Bq/g]) (SRM 4324) and 238U, "natural uranium"
(6.7 nCi/g [250 Bq/g]) (SRM 4321B) (NIST 1995). Standard Reference Materials of human lung (SRM 4351) and human liver (SRM 4352) are also available from NIST.
6.3 ADEQUACY OF THE DATABASE
Section 104(i)(5) of CERCLA, directs the Administrator of ATSDR (in consultation with the Administrator of EPA and agencies and programs of the Public Health Service) to assess whether adequate information on the health effects of uranium is available. Where adequate information is not available, ATSDR, in conjunction with the NTP, is required to assure the initiation of a program of research designed to determine the health effects (and techniques for developing methods to determine such health effects) of uranium.
The following categories of possible data needs have been identified by a joint team of scientists from ATSDR, NTP, and EPA. They are defined as substance-specific informational needs that, if met, would reduce or eliminate the uncertainties of human health assessment. In the future, the identified data needs will be evaluated and prioritized, and a substance-specific research agenda will be proposed.
6.3.1 Identification of Data Needs
Methods for Determining Biomarkers of Exposure and Effect. Analytical methods with satisfactory sensitivity and precision are available to determine the levels of uranium in human tissues and body fluids. However, improved methods are needed to assess the biological effects of uranium in tissues.
Uranium is in essentially all food, water, and air, so everyone is exposed to some levels. In a study reported by NIOSH (Thun et al. 1981, 1985), enhanced levels of β2-microglobulin levels were observed in the urine of uranium workers. It was postulated that enhanced excretion of β2-microglobulin might be used as an indication of uranium exposure; however, Thun et al. (1981, 1985) were unable to establish a dose response correlation between level of exposure and excretion of the β2-microglobulin. Limson-Zamora et al. (1996) identified changes in several potential biomarkers of effect following exposure to uranium, in which each
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individual biomarker could be affected by a range of chemicals, but the results suggested that it may be possible to identify a series of biomarkers whose combined responses could serve as a single uranium-specific biomarker of effect. Development of new or combination biomarkers for high uranium exposures would be useful.
Methods for Determining Parent Compounds and Degradation Products in Environmental Media. Analytical methods with the required sensitivity and accuracy are available for quantification of uranium, both total and isotopic, in environmental matrices (Table 6-2). Knowledge of the levels of uranium in various environmental media, along with the appropriate modeling (see Chapters 2 and 4), can be used to evaluate potential human exposures through inhalation and ingestion pathways.
Whether in the environment or in the human body, uranium will undergo radioactive decay to form a series of radioactive nuclides that end in a stable isotope of lead (see Chapter 3). Examples of these include radioactive isotopes of the elements thorium, radium, radon, polonium, and lead. Analytical methods with the required sensitivity and accuracy are also available for quantification of these elements in the environment where large sample are normally available (EPA 1980, 1984), but not necessarily for the levels from the decay of uranium in the body. More sensitive analytical methods are needed for accurately measuring very low levels of these radionuclides.
6.3.2 Ongoing Studies
The Federal Research in Progress (FEDRIP) database lists ongoing studies investigating new methods for detection and speciation of uranium (FEDRIP 1999). These are shown in Table 6-4.
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The international, national, and state regulations and guidelines regarding uranium in air, water, and other media are summarized in Table 7-1.
ATSDR has derived an MRL of 8.0×10-3 mg U/m3 for intermediate-duration inhalation exposure to insoluble compounds of uranium based on a NOAEL of 1.1 mg U/m3 for renal effects in dogs (Rothstein 1949b).
ATSDR has derived an MRL of 4.0×10-4 mg U/m3 for intermediate-duration inhalation exposure to soluble compounds of uranium based on a LOAEL of 0.15 mg U/m3 for renal effects in dogs (Rothstein 1949a).
ATSDR has derived an MRL of 3.0×10-4 mg U/m3 for chronic-duration inhalation exposure (365 days or more) to soluble compounds of uranium based on a NOAEL of 0.05 mg U/m3 for renal effects in dogs (Stokinger et al. 1953).
ATSDR has derived an MRL of 2.0×10-3 mg/kg/day has been derived for intermediate-duration oral exposure (and is protective for chronic-duration oral exposure) to soluble compounds of uranium based on a LOAEL of 0.05 mg U/kg/day d for renal effects in rabbits (Gilman et al. 1998b).
According to the EPA’s Integrated Risk Information System (IRIS), neither a reference dose (RfD) nor a reference concentration (RfC) are available for uranium (IRIS 1997).
The International Agency for Research on Cancer, the U.S. Department of Human and Health Services, and the National Toxicology Program have not classified uranium as to its carcinogenicity. According to the Integrated Risk Information System (IRIS) database, the U.S. Environmental Protection Agency (EPA) withdrew its carcinogenic assessment of uranium in 1993 and has not completed its evaluation and determination of the evidence of uranium’s human carcinogenicity potential (IRIS 1997).
Radiation protection recommendations for radiation workers and members of the public are provided by the International Commission on Radiological Protection (ICRP) (ICRP 1977, 1991) and the National Council on Radiation Protection and Measurements (NCRP) (NCRP 1987, 1993). These recommendations are not
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regulations, but they provide the scientific basis for the development of regulations by federal agencies, such as EPA, the U.S. Nuclear Regulatory Commission (USNRC), and the U.S. Department of Energy (DOE), as well as by individual states.
The EPA is responsible for federal radiation protection guidance (EPA 1988c), "generally applicable" environmental radiation standards (40 CFR 190), and regulations to implement specific statutory requirements, such as Safe Drinking Water Act (40 CFR 141). The USNRC’s regulations apply to source materials and special nuclear material, such as enriched uranium and plutonium; the utilization of special nuclear material, such as the operation of nuclear reactors; and the use of by-product materials, which include wastes produced in the processing of uranium or thorium and materials made radioactive in the utilization of special nuclear material (USNRC 1997a). The DOE has issued regulations applicable to its facilities (DOE 1993a).
States are free to regulate radioactive materials and other sources of radiation that the Atomic Energy Act does not give the USNRC authority to regulate. This includes sources of natural radioactivity, such as, uranium and radium, and radiation producing machines, such as x ray machines. Section 274 of the Atomic Energy Act (AEA) of 1954, as amended provides that states (and U.S. territories) may enter into an agreement with the USNRC to regulate by-product materials, source materials, and special nuclear materials (USNRC 1969). The relinquishes to these “agreement states” the majority of its regulatory authority over source, by-product, and special nuclear material in quantities not sufficient to form a critical mass. However, the USNRC retains its authority to regulate the construction and operation of production facilities (nuclear reactors used for production and separation of plutonium or 233U or fuel reprocess plants) and utilization facilities (nuclear reactors used for production of power, medical therapy, research and testing); the import of by-product, source, or special nuclear materials; and the disposal of regulated materials into the ocean or otherwise (USNRC 1969). Currently there are 30 “agreements states” and 17 “non-agreement states.” The governors of five states (Minnesota, Ohio, Oklahoma, Pennsylvania, and Wisconsin) have submitted letters of intent for their states to become agreement states (ORNL 1998). The regulations established by agreement states must be "compatible" with the USNRC's regulations, which require that the states' regulations be at least as strict as the USNRC's regulations. The responsibilities of agreement states also include the regulation of low-level radioactive wastes, which contain by-product materials. In nonagreement states, the USNRC still handles all of the inspection, enforcement, and licensing responsibilities. Figure 7-1 shows the agreement, non-agreement, and intending states.
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Current federal and state regulations limit radiation workers’ doses to a total effective dose equivalent (TEDE) of 5 rem/year and a committed dose equivalent to any organ, other than the lens of the eye, of 50 rem/year (EPA 1988c; USNRC 1995a). These limits apply to the sum of external and internal doses. The limits are upper limits, and an important philosophy in radiation protection is to keep radiation doses as low as reasonably achievable (ALARA).
For the control of internal doses, annual limits of intake (ALI) and derived air concentrations (DAC) have been determined. ALIs and DACs in EPA guidance and the USNRC and DOE regulations are based on the recommendations of the ICRP (ICRP 1979). Values of the ALIs and DACs for uranium isotopes are presented in Table 7-1. These values are for soluble, Class D (Days) material, which has a half-time for clearance from the pulmonary region of the lung of less than 10 days. Values of ALIs and DACs for Class W (Weeks) and Class Y (Years) uranium are available in Appendix B to 10 CFR 20 (USNRC 1993f).
The ALI is the activity of a radionuclide that can be taken into the body in a year, by inhalation or ingestion, without exceeding a committed effective dose equivalent (CEDE) of 5 rem/year or a committed dose equivalent to any organ of 50 rem/year, whichever is more limiting. The total effective dose equivalent TEDE is the sum of the CEDE and any penetrating external dose (10 CFR 20). If any external dose is present the ALI must be reduced by a proportional amount to ensure that the dose limits are not exceeded. For example, if a worker received an external dose of 1 rem/year, the ALI would have to be reduced by 20% to ensure that the TEDE did not exceed 5 rem/year.
The DAC is simply the inhalation ALI divided by the volume of air that a worker is assumed to breathe in a year (2,400 m3). Thus, if the average air concentration is controlled so as not to exceed the DAC, a worker will not take in more than an ALI, and the worker’s dose will not exceed 5 rem CEDE or 50 rem committed dose equivalent to any organ (ICRP 1977).
Uranium is unusual among the elements because it presents both chemical and radiological hazards. For soluble uranium, with an 235U enrichment no greater than 5%, limits on intakes and air concentrations for radiation workers are based on the chemical toxicity of uranium since it is more limiting than the radiological hazard. For this case, the USNRC’s limit for a 40-hour workweek is 0.2 mg uranium per cubic meter of air average (USNRC 1993f).
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Between June 27, 1974 and January 18, 1989, the Occupational Safety and Health Administration (OSHA) promulgated protective, permissible exposure limits (PELs) for more than 400 toxic substances (OSHA 1993). The OSHA PELs were established to protect employees against adverse health effects which could result from exposure to hazardous substances found in the workplace. An employer must ensure that an employee’s exposure to a toxic substance in any 8-hour work shift of a 40-hour week does not exceed the 8-hour time-weighted average (TWA) established for the substance (OSHA 1993). On January 18, 1989 OSHA promulgated more protective PELs for approximately 376 toxic substances. In July 1992, the 11th Circuit Court Appeals rescinded the 1989 promulgation. On March 23, 1993, OSHA resumed enforcing the air contaminant exposure limits that were in effect prior to the issuance of the new 1989 limits (i.e., OSHA 1974 PELs). On June 30, 1993 OSHA published in the Federal Register a final rule announcing the revocation of the 1989 exposure limits. Current OSHA general industry standards specify that an employer must use engineering and work practice controls, if feasible, to reduce exposures to or below an 8-hour TWA of 0.05 mg per cubic meter of soluble uranium and 0.25 mg/m3 for insoluble uranium (OSHA 1993, 1997c). OSHA standards for construction workers indicate that exposures to uranium in gases, vapors, fumes, dust, and mist, through inhalation, ingestion, skin absorption, or contact at concentrations above those specified in the ACGIH “Threshold Limit Values of Airborne Contaminants for 1970" should be avoided. The “Threshold Limit Values of Airborne Contaminants for Construction” are codified at 29 CFR 1926, and indicate a limit of exposure at 0.2 mg/m3 for soluble and insoluble uranium (OSHA 1997a). The same limits of exposure are codified at 29 CFR 1915 for shipyard personnel (OSHA 1997b).
Recent reports of the ICRP (ICRP 1991) and the NCRP (NCRP 1993) contain recommendations for lower worker dose limits. The ICRP recommends a limit on total effective dose of 2 rem/year averaged over 5 years, with the additional provision that the dose not exceed 5 rem in any single year. The NCRP's recommendations are that a worker’s total accumulated dose should not exceed his or her age in years time 1 rem, and that the dose should not exceed 5 rem in any single year. These recommendations have not yet been incorporated into U.S. regulations.
Regulations for the general public are based on an annual TEDE of 0.1 rem/year, with provisions for a limit of 0.5 rem/year under special circumstances (USNRC 1997b). Considering the lower limit for members of the public and their potential continuous exposure, the limits on air concentrations of radionuclides for the public are two orders of magnitude lower than the DACs for radiation workers. Regulations for specific applications limit the dose to the public to values 90% in a gaseous diffusion process based on the different thermal velocities of the constituents of natural uranium (234U, 235U, 238U) in the molecular form UF6.
EPA Health Advisory—An estimate of acceptable drinking water levels for a chemical substance based on health effects information. A health advisory is not a legally enforceable federal standard, but serves as technical guidance to assist federal, state, and local officials.
Equilibrium, Radioactive—In a radioactive series, the state which prevails when the ratios between the activities of two or more successive members of the series remains constant.
Secular Equilibrium—If a parent element has a very much longer half-life than the daughters (so there is not appreciable change in its amount in the time interval required for later products to attain equilibrium) then, after equilibrium is reached, equal numbers of atoms of all members of the series disintegrate in unit time. This condition is never exactly attained, but is essentially established in such a case as 226Ra and its transformation series to stable 206Pb. The half-life of 226Ra is about 1,600 years; of 222Rn, approximately 3.82 days, and of each of the subsequent members, a few minutes. After about a month, essentially the equilibrium amount of radon is present; then (and for a long time) all members of the series disintegrate the same number of atoms per unit time. At this time, the activity of the daughter is equal to the activity of the parent.
Transient Equilibrium—If the half-life of the parent is short enough so the quantity present decreases appreciably during the period under consideration, but is still longer than that of successive members of the series, a stage of equilibrium will be reached after which all members of the series decrease in activity exponentially with the period of the parent. At this time, the ratio of the parent activity to the daughter activity is constant.
Equilibrium, Electron—The condition in a radiation field where the energy of the electrons entering a volume equals the energy of the electrons leaving that volume.
Equilibrium Fraction (F)—In radon-radon daughter equilibrium, the parents and daughters have equal radioactivity, that is, as many decay into a specific nuclide as decay out. However, if fresh radon is continually entering a volume of air or if daughters are lost by processes other than radioactive decay, e.g., plate out or migration out of the volume, a disequilibrium develops. The equilibrium fraction is a measure of the degree of equilibrium/disequilibrium. The equilibrium fraction is used to estimate working levels based on measurement of radon only. For radon, 1 working-level concentration is defined at 100 pCi of radon in equilibrium with its 4 successive progeny in 1 liter of air. Thus, 100 pCi/L radon at 50% equilibrium is 0.5 WL.
Excitation—The addition of energy to a system, thereby transferring it from its ground state to an excited state. Excitation of a nucleus, an atom, or a molecule can result from absorption of photons or from inelastic collisions with other particles. The excited state of an atom is an unstable or metastable state and will return to ground state by radiation of the excess energy.
Exposure (Chemical)—Contact of an organism with a chemical or physical agent. Exposure is quantified as the amount of the agent available at the exchange boundaries of the organism (e.g., skin, lungs, gut) and available for absorption.
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Exposure (Radiation)—Being exposed to ionizing radiation or to a radioactive material. A measure of the ionization produced in air by x or gamma radiation; the sum of the electric charges on all ions of one sign produced in air when all electrons liberated by photons in a volume of air are completely stopped in air (dQ), divided by the mass of the air in the volume (dm). The unit of exposure in air is the roentgen, or coulomb per kilogram (SI units). One roentgen is equal to 2.58×10-4 coulomb per kilogram (C/kg).
Fission, Nuclear—A nuclear transformation characterized by the splitting of a nucleus into at least two other nuclei and several neutrons, and is accompanied by the release of a relatively large amount of energy.
Gamma Ray, Penetrating—Short wavelength electromagnetic radiation of nuclear origin.
Genetic Effect of Radiation—Inheritable change, chiefly mutations, produced by the absorption of ionizing radiation by germ cells. Genetic effects have not been observed in any human population exposed at any dose level.
Gray (Gy)—SI unit of absorbed dose, 1 J/kg. One gray equals 100 rad (see Units).
Half-life, Radioactive—Time required for a radioactive substance to lose 50% of its activity by decay. Each radio-nuclide has a unique physical half-life. Known also as physical half-time and symbolized as Tr or Trad.
Half-life, Effective—See Half-Time, Effective.
Half-time, Biological—Time required for an organ, tissue, or the whole body to eliminate one-half of any absorbed substance by regular processes of elimination. This is the same for both stable and radioactive isotopes of a particular element, and is sometimes referred to as half-time, symbolized as tbiol or Tb.
Half-time, Effective—Time required for a radioactive element in an organ, tissue, or the whole body to be diminished 50% as a result of the combined action of radioactive decay and biological elimination, symbolized as Te or Teff.
Effective Half&time ‘ Biological half&time x Radioactive half&lifeBiological half&time % Radioactive half&life
Immediately Dangerous to Life or Health (IDLH)—The maximum environmental concentration of a contaminant from which one could escape within 30 minutes without any escape-impairing symptoms or irreversible health effects.
Immunologic Toxicity—The occurrence of adverse effects on the immune system that may result from exposure to environmental agents such as chemicals.
In Vitro—Isolated from the living organism and artificially maintained, as in a test tube. Literally, “in glass.”
In Vivo—Occurring within the living organism. Literally, “in life.”
Intensity—Amount of energy per unit time passing through a unit area perpendicular to the line of propagation at the point in question.
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Intermediate Exposure—Exposure to a chemical for a duration of 15–364 days, as specified in the Toxicological Profiles.
Internal Conversion—Process in which a gamma ray knocks an electron out of the same atom from which the gamma ray was emitted. The ratio of the number of internal conversion electrons to the number of gamma quanta emitted in the de-excitation of the nucleus is called the “conversion ratio.”
Ion—Atomic particle, atom or chemical radical bearing a net electrical charge, either negative or positive.
Ion Pair—Two particles of opposite charge, usually referring to the electron and positive atomic or molecular residue resulting after the interaction of ionizing radiation with the orbital electrons of atoms.
Ionization—The process by which a neutral atom or molecule acquires a positive or negative charge.
Primary Ionization—(1) In collision theory: the ionization produced by the primary particles as contrasted to the “total ionization” which includes the “secondary ionization” produced by delta rays. (2) In counter tubes: the total ionization produced by incident radiation without gas amplification.
Specific Ionization—Number of ion pairs per unit length of path of ionizing radiation in a medium; e.g., per centimeter of air or per micrometer of tissue.
Total Ionization—The total electric charge of one sign on the ions produced by radiation in the process of losing its kinetic energy. For a given gas, the total ionization is closely proportional to the initial ionization and is nearly independent of the nature of the ionizing radiation. It is frequently used as a measure of absorption of radiation energy.
Ionization Density—Number of ion pairs per unit volume.
Ionization Path (Track)—The trail of ion pairs produced by an ionizing particle in its passage through matter.
Ionizing Radiation—Any radiation capable of knocking electrons out of atoms and producing ions. Examples: alpha, beta, gamma and x rays, and neutrons.
Isobars—Nuclides having the same mass number but different atomic numbers.
Isomers—Nuclides having the same number of neutrons and protons but capable of existing, for a measurable time, in different quantum states with different energies and radioactive properties. Commonly the isomer of higher energy decays to one with lower energy by the process of isomeric transition.
Isotopes—Nuclides having the same number of protons in their nuclei, and hence the same atomic number, but differing in the number of neutrons, and therefore in the mass number. Identical chemical properties exist in isotopes of a particular element. The term should not be used as a synonym for nuclide because isotopes refer specifically to different nuclei of the same element.
Stable Isotope—A nonradioactive isotope of an element.
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9. GLOSSARY
Kerma (k)—A measure of the kinetic energy transferred from gamma rays or neutrons to a unit mass of absorbing medium in the initial collision between the radiation and the absorber atoms. The SI unit is J/kg. The special name of this unit is the rad (traditional system of units) or Gray (SI).
Joule—The S.I. unit for work and energy. It is equal to the work done by raising a mass of one newton through a distance of one meter (J = Nm), which corresponds to about 0.7 ft-pound.
Labeled Compound—A compound containing one or more radioactive atoms intentionally added to its structure. By observations of radioactivity or isotopic composition, this compound or its fragments may be followed through physical, chemical, or biological processes.
Late Effects (of radiation exposure)—Effects which appear 60 days or more following an acute exposure.
LD50/30—The dose of a chemical or radiation expected to cause 50% mortality in those exposed within 30 days. For radiation, this is about 350 rad (3.5 gray) received by humans over a short period of time.
Lethal Concentration(LO) (LCLO)—The lowest concentration of a chemical in air that has been reported to have caused death in humans or animals.
Lethal Concentration(50) (LC50)—A calculated concentration of a chemical in air to which exposure for a specific length of time is expected to cause death in 50% of a defined experimental animal population within a specified time, usually 30 days.
Lethal Dose(L0) (LDL0)—The lowest dose of a chemical introduced by a route other than inhalation that is expected to have caused death in humans or animals within a specified time, usually 30 days.
Lethal Dose(50) (LD50)—The dose of a chemical which has been calculated to cause death in 50% of a defined experimental animal population.
Lethal Time(50) (LT50)—A calculated period of time within which a specific concentration of a chemical is expected to cause death in 50% of a defined experimental animal population.
Linear Energy Transfer (LET)—A measure of the energy that a charged particle transfers to a material per unit path length.
Low-LET—Energy transfer characteristic of light charged particles such as electrons produced by x and gamma rays where the distance between ionizing events is large on the scale of a cellular nucleus.
High-LET—Energy transfer characteristic of heavy charged particles such as protons and alpha particles where the distance between ionizing events is small on the scale of a cellular nucleus.
Average LET—The energy of a charged particle divided by the length of the path over which it deposits all its energy in a material.
Lowest-Observed-Adverse-Effect Level (LOAEL)—The lowest dose of chemical in a study, or group of studies, that produces statistically or biologically significant increases in frequency or severity of adverse effects between the exposed population and its appropriate control.
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9. GLOSSARY
Lung Clearance Class (fast, F; medium, M; slow, S)—A classification scheme for inhaled material according to its rate of clearance from the pulmonary region of the lungs to the blood and the gastrointestinal tract.
Malformations—Permanent structural changes that may adversely affect survival, development, or function.
Mass Numbers (A)—The number of nucleons (protons and neutrons) in the nucleus of an atom.
Minimal Risk Level—An estimate of daily human exposure to a substance that is likely to be without an appreciable risk of adverse noncancerous effects over a specified duration of exposure.
Mutagen—A substance that causes changes (mutations) in the genetic material in a cell. Mutations can lead to birth defects, miscarriages, or cancer.
Neurotoxicity—The occurrence of adverse effects on the nervous system following exposure to a substance.
Neutrino (<)—A neutral particle of infinitesimally small rest mass emitted during beta plus or beta minus decay. This particle accounts for conservation of energy in beta plus and beta minus decays. It plays no role in damage from radiation.
No-Observed-Adverse-Effect Level (NOAEL)—The dose of a substance at which there were no statistically or biologically significant increases in frequency or severity of adverse effects seen between the exposed population and its appropriate control. Effects may be produced at this dose, but they are not considered to be adverse.
Nuclear reactor—A power plant that heats water by using nuclear reactions instead of burning coal, oil, or natural gas. All of these sources of energy simply heat water and use the steam which is produced to turn turbines that make electricity or propel a ship.
Nucleon—Common name for a constituent particle of the nucleus. Applied to a proton or neutron.
Nuclide—A species of atom characterized by the constitution of its nucleus. The nuclear constitution is specified by the number of protons (Z), number of neutrons (N), and energy content; or, alternatively, by the atomic number (Z), mass number A'(N+Z), and atomic mass. To be regarded as a distinct nuclide, the atom must be capable of existing for a measurable time. Thus, nuclear isomers are separate nuclides, whereas promptly decaying excited nuclear states and unstable intermediates in nuclear reactions are not so considered.
Octanol-Water Partition Coefficient (Kow)—The equilibrium ratio of the concentrations of a chemical in n-octanol and water, in dilute solution.
Pair Production—An absorption process for x- and gamma radiation in which the incident photon is absorbed in the vicinity of the nucleus of the absorbing atom, with subsequent production of an electron and positron pair (see annihilation). This reaction can only occur for incident photon energies exceeding 1.02 MeV.
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9. GLOSSARY
Parent—A radionuclide which, upon disintegration, yields a new nuclide, either directly or as a later member of a radioactive series.
Permissible Exposure Limit (PEL)—A maximum allowable atmospheric level of a substance in workplace air averaged over an 8-hour shift.
Photon—A quantum of electromagnetic energy (E) whose value is the product of its frequency (ν) in hertz and Planck's constant (h). The equation is: E = hν.
Photoelectric Effect—An attenuation process observed for x and gamma radiation in which an incident photon interacts with a tightly bound inner orbital electron of an atom delivering all of its energy to knock the electron out of the atom. The incident photon disappears in the process.
Positron—A positively charged electron.
Potential, Ionization—The energy expressed as electron volts (eV) necessary to separate one electron from an atom, resulting in the formation of an ion pair.
Power, Stopping—A measure of the ability of a material to absorb energy from an ionizing particle passing through it; the greater the stopping power, the greater the energy absorbing ability (see Linear Energy Transfer).
Progeny—The decay product or products resulting after a radioactive decay or a series of radioactive decays. The progeny can also be radioactive, and the chain continues until a stable nuclide is formed.
Proton—Elementary nuclear particle with a positive electric charge equal numerically to the charge of the electron and a rest mass of 1.007 mass units.
Quality—A term describing the distribution of the energy deposited by a particle along its track; radiations that produce different densities of ionization per unit intensity are said to have different "qualities."
Quality Factor (Q)—The linear-energy-transfer-dependent factor by which absorbed doses are multiplied to obtain (for radiation protection purposes) a quantity that expresses – on a common scale for all ionizing radiation – the approximate biological effectiveness of the absorbed dose.
Rad—The unit of absorbed dose equal to 100 ergs per gram, or 0.01 joule per kilogram (0.01 Gy) in any medium (see Absorbed Dose).
Radiation—The emission and propagation of energy through space or through a material medium in the form of waves (e.g., the emission and propagation of electromagnetic waves, or of sound and elastic waves). The term radiation or radiant energy, when unqualified, usually refers to electromagnetic radiation. Such radiation commonly is classified according to frequency, as microwaves, infrared, visible (light), ultraviolet, and x and gamma rays (see Photon.) and, by extension, corpuscular emission, such as alpha and beta radiation, neutrons, or rays of mixed or unknown type, as cosmic radiation.
Radiation, Annihilation—Photons produced when an electron and a positron unite and cease to exist. The annihilation of a positron-electron pair results in the production of two photons, each of 0.51 MeV energy.
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9. GLOSSARY
Radiation, Background—See Background Radiation.
Radiation, Characteristic (Discrete)—Radiation originating from an excited atom after removal of an electron from an atom. The wavelength of the emitted radiation is specific, depending only on the element and particular energy levels involved.
Radiation, External—Radiation from a source outside the body.
Radiation, Internal—Radiation from a source within the body (as a result of deposition of radionuclides in body tissues).
Radiation, Ionizing—Any electromagnetic or particulate radiation capable of producing ions, directly or indirectly, in its passage through matter (see Radiation).
Radiation, Monoenergetic—Radiation of a given type in which all particles or photons originate with and have the same energy.
Radiation, Scattered—Radiation which during its passage through a substance, has been deviated in direction. It may also have been modified by a decrease in energy.
Radiation, Secondary—A particle or ray that is produced when the primary radiation interacts with a material, and which has sufficient energy to produce its own ionization, such as bremsstrahlung or electrons knocked from atomic orbitals with enough energy to then produce ionization (see Delta Rays).
Radioactive Material—Material containing radioactive atoms.
Radioactivity—Spontaneous nuclear transformations that result in the formation of new elements. These transformations are accomplished by emission of alpha or beta particles from the nucleus or by the capture of an orbital electron. Each of these reactions may or may not be accompanied by a gamma photon.
Radioactivity, Artificial—Man-made radioactivity produced by particle bombardment or nuclear fission, as opposed to naturally occurring radioactivity.
Radioactivity, Induced—Radioactivity produced in a substance after bombardment with neutrons or other particles. The resulting activity is "natural radioactivity" if formed by nuclear reactions occurring in nature and "artificial radioactivity" if the reactions are caused by man.
Radioactivity, Natural—The property of radioactivity exhibited by more than 50 naturally occurring radionuclides.
Radioisotope—An unstable or radioactive isotope of an element that decays or disintegrates spontaneously, emitting radiation. Approximately 5,000 natural and artificial radioisotopes have been identified.
Radionuclide—Any radioactive isotope of any element.
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9. GLOSSARY
Radiosensitivity—Relative susceptibility of cells, tissues, organs, organisms, or any living substance to the injurious action of radiation. Radiosensitivity and its antonym, radioresistance, are currently used in a comparative sense, rather than in an absolute one.
Reference Dose (RfD)—An estimate (with uncertainty spanning perhaps an order of magnitude) of the daily exposure of the human population to a potential hazard that is likely to be without risk of deleterious effects during a lifetime. The RfD is operationally derived from the NOAEL (from animal and human studies) by a consistent application of uncertainty factors that reflect various types of data used to estimate RfDs and an additional modifying factor, which is based on a professional judgment of the entire database on the chemical. The RfDs are not applicable to non-threshold effects such as cancer.
Relative Biological Effectiveness (RBE)—The RBE is a factor used to compare the biological effectiveness of absorbed radiation doses (i.e., rad) due to different types of ionizing radiation. More specifically, it is the experimentally determined ratio of an absorbed dose of a radiation in question to the absorbed dose of a reference radiation (typically 60Co gamma rays or 200 keV x rays) required to produce an identical biological effect in a particular experimental organism or tissue (see Quality Factor).
Rem—A unit of dose equivalent that is used in the regulatory, administrative, and engineering design aspects of radiation safety practice. The dose equivalent in rem is numerically equal to the absorbed dose in rad multiplied by the quality factor (1 rem is equal to 0.01 sievert).
Reportable Quantity (RQ)—The quantity of a hazardous substance that is considered reportable under CERCLA. Reportable quantities are (1) 1 pound or greater or (2) for selected substances, an amount established by regulation either under CERCLA or under Sect. 311 of the Clean Water Act. Quantities are measured over a 24-hour period.
Reproductive Toxicity—The occurrence of adverse effects on the reproductive system that may result from exposure to a chemical. The toxicity may be directed to the reproductive organs and/or the related endocrine system. The manifestation of such toxicity may be noted as alterations in sexual behavior, fertility, pregnancy outcomes, or modifications in other functions that are dependent on the integrity of this system.
Roentgen (R)—A unit of exposure (in air) to ionizing radiation. It is the amount of x or gamma rays required to produce ions carrying 1 electrostatic unit of electrical charge in 1 cubic centimeter of dry air under standard conditions. Named after William Roentgen, a German scientist who discovered x rays in 1895.
Self-Absorption—Absorption of radiation (emitted by radioactive atoms) by the material in which the atoms are located; in particular, the absorption of radiation within a sample being assayed.
Short-Term Exposure Limit (STEL)—The maximum concentration to which workers can be exposed for up to 15 min continually. No more than four excursions are allowed per day, and there must be at least 60 min between exposure periods. The daily TLV-TWA may not be exceeded.
SI Units—The International System of Units as defined by the General Conference of Weights and Measures in 1960. These units are generally based on the meter/kilogram/second units, with special quantities for radiation including the becquerel, gray, and sievert.
Sickness, Acute Radiation (Syndrome)—The complex symptoms and signs characterizing the condition resulting from excessive exposure of the whole body (or large part) to ionizing radiation. The earliest of
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9. GLOSSARY
these symptoms are nausea, fatigue, vomiting, and diarrhea, and may be followed by loss of hair (epilation), hemorrhage, inflammation of the mouth and throat, and general loss of energy. In severe cases, where the radiation dose is relatively high (over several hundred rad or several gray), death may occur within two to four weeks. Those who survive six weeks after exposure of a single high dose of radiation may generally be expected to recover.
Sievert (Sv)—The SI unit of any of the quantities expressed as dose equivalent. The dose equivalent in sieverts is equal to the absorbed dose, in gray, multiplied by the quality factor (1 sievert equals 100 rem).
Specific-activity—Radioactivity per unit mass of material containing a radionuclide, expressed, for example, as Ci/gram or Bq/gram.
Specific Energy—The actual energy per unit mass deposited per unit volume in a small target, such as the cell or cell nucleus, as the result of one or more energy-depositing events. This is a stochastic quantity as opposed to the average value over a large number of instance (i.e., the absorbed dose).
Standard Mortality Ratio (SMR)—A ratio of the observed number of deaths and the expected number of deaths in a specific standard population.
Stopping Power—The average rate of energy loss of a charged particle per unit thickness of a material or per unit mass of material traversed.
Surface-seeking Radionuclide—A bone-seeking internal emitter that is deposited and remains on the bone surface for a long period of time, although it may eventually diffuse into the bone mineral. This contrasts with a volume seeker, which deposits more uniformly throughout the bone volume.
Target Organ Toxicity—This term covers a broad range of adverse effects on target organs or physiological systems (e.g., renal, cardiovascular) extending from those arising through a single limited exposure to those assumed over a lifetime of exposure to a chemical.
Target Theory (Hit Theory)—A theory explaining some biological effects of radiation on the basis that ionization, occurring in a discrete volume (the target) within the cell, directly causes a lesion which subsequently results in a physiological response to the damage at that location. One, two, or more "hits" (ionizing events within the target) may be necessary to elicit the response.
Teratogen—A chemical that causes birth defects.
Threshold Limit Value (TLV)—The maximum concentration of a substance to which most workers can be exposed without adverse effect. TLV is a term used exclusively by the ACGIH. Other terms used to express the same concept are the MAC (Maximum Allowable Concentration) and PEL (Permissible Exposure Limits).
Time-Weighted Average (TWA)—An allowable exposure concentration averaged over a normal 8-hour workday or 40-hour workweek.
Toxic Dose (TD50)—A calculated dose of a chemical, introduced by a route other than inhalation, which is expected to cause a specific toxic effect in 50% of a defined experimental animal population.
Toxicosis —A diseased condition resulting from poisoning.
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9. GLOSSARY
Transformation, Nuclear—The process by which a nuclide is transformed into a different nuclide by absorbing or emitting particulate or electromagnetic radiation.
Transition, Isomeric—The process by which a nuclide decays to an isomeric nuclide (i.e., one of the same mass number and atomic number) of lower quantum energy. Isomeric transitions (often abbreviated I.T.) proceed by gamma ray and/or internal conversion electron emission.
Tritium—The hydrogen isotope with one proton and two neutrons in the nucleus (Symbol: 3H). It is radioactive and has a physical half-life of 12.3 years.
Unattached Fraction—That fraction of the radon daughters, usually 218Po and 214Po, which has not yet attached to a dust particle or to water vapor. As a free atom, it has a high probability of being exhaled and not retained within the lung. It is the attached fraction which is primarily retained.
Uncertainty Factor (UF)—A factor used in operationally deriving the RfD from experimental data. UFs are intended to account for (1) the variation in sensitivity among the members of the human population, (2) the uncertainty in extrapolating animal data to the case of human, (3) the uncertainty in extrapolating from data obtained in a study that is of less than lifetime exposure, and (4) the uncertainty in using LOAEL data rather than NOAEL data. Usually each of these factors is set equal to 10.
Units, Radiological—
Units Equivalents
Becquerel* (Bq) 1 disintegration per second = 2.7×10-11 Ci
Curie (Ci) 3.7×1010 disintegrations per second = 3.7×1010 Bq
Gray* (Gy) 1 J/kg = 100 rad
Rad (rad) 100 erg/g = 0.01 Gy
Rem (rem) 0.01 sievert
Sievert* (Sv) 100 rem

*International Units, designated (SI)
Working Level (WL)—Any combination of short-lived radon daughters in 1 liter of air that will result in the ultimate emission of 1.3×105 MeV of potential alpha energy.
Working Level Month (WLM)—A unit of exposure to radon daughters corresponding to the product of the radon daughter concentration in Working Level (WL) and the exposure time in nominal months (1 nominal month = 170 hours). Inhalation of air with a concentration of 1 WL of radon daughters for 170 working hours results in an exposure of 1 WLM.
X rays—Penetrating electromagnetic radiations whose wave lengths are very much shorter than those of visible light. They are usually produced by bombarding a metallic target with fast electrons in a high vacuum. X rays (called characteristic x rays) are also produced when an orbital electron falls from a high energy level to a low energy level.
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9. GLOSSARY
Zero-Threshold Linear Hypothesis—The assumption that a dose-response curve derived from data in the high dose and high dose-rate ranges may be extrapolated through the low dose and low dose range to zero, implying that, theoretically, any amount of radiation will cause some damage.
URANIUM A-1

APPENDIX AATSDR MINIMAL RISK LEVELS AND WORKSHEETS
The Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA) [42 U.S.C. 9601 et seq.], as amended by the Superfund Amendments and Reauthorization Act (SARA) [Pub. L. 99–499], requires that the Agency for Toxic Substances and Disease Registry (ATSDR) develop jointly with the U.S. Environmental Protection Agency (EPA), in order of priority, a list of hazardous substances most commonly found at facilities on the CERCLA National Priorities List (NPL); prepare toxicological profiles for each substance included on the priority list of hazardous substances; and assure the initiation of a research program to fill identified data needs associated with the substances.
The toxicological profiles include an examination, summary, and interpretation of available toxicological information and epidemiologic evaluations of a hazardous substance. During the development of toxicological profiles, Minimal Risk Levels (MRLs) are derived when reliable and sufficient data exist to identify the target organ(s) of effect or the most sensitive health effect(s) for a specific duration for a given route of exposure. An MRL is an estimate of the daily human exposure to a hazardous substance that is likely to be without appreciable risk of adverse noncancer health effects over a specified duration of exposure. MRLs are based on noncancer health effects only and are not based on a consideration of cancer effects. These substance-specific estimates, which are intended to serve as screening levels, are used by ATSDR health assessors to identify contaminants and potential health effects that may be of concern at hazardous waste sites. It is important to note that MRLs are not intended to define clean-up or action levels.
MRLs are derived for hazardous substances using the no-observed-adverse-effect-level/uncertainty factor approach. They are below levels that might cause adverse health effects in the people most sensitive to such effects. MRLs are derived for acute (1–14 days), intermediate (15–364 days), and chronic (365 days and longer) durations and for the oral, inhalation, and external routes of exposure. Currently, MRLs for the dermal route of exposure are not derived because ATSDR has not yet identified a method suitable for this route of exposure. MRLs are generally based on the most sensitive end point considered to be a relevance to humans. Serious health effects (such as irreparable damage to the liver or kidneys, or birth defects) are not used as a basis for establishing MRLs. Exposure to a level above the MRL does not mean that adverse health effects will occur.
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APPENDIX A
MRLs are intended only to serve as a screening tool to help public health professionals decide where to look more closely. They may also be viewed as a mechanism to identify those hazardous waste sites that are not expected to cause adverse health effects. Most MRLs contain a degree of uncertainty because of the lack of precise toxicological information on the people who might be most sensitive (e.g., infants, elderly, nutritionally or immunologically compromised) to the effects of hazardous substances. ATSDR uses a conservative (i.e., protective) approach to address this uncertainty consistent with the public health principle of prevention. Although human data are preferred, MRLs often must be based on animal studies because relevant human studies are lacking. In the absence of evidence to the contrary, ATSDR assumes that humans are more sensitive to the effects of hazardous substance than animals and that certain persons may be particularly sensitive.
Proposed MRLs undergo a rigorous review process: Health Effects/MRL Workgroup reviews within the Division of Toxicology, expert panel peer reviews, and agencywide MRL Workgroup reviews, with participation from other federal agencies and comments from the public. They are subject to change as new information becomes available concomitant with updating the toxicological profiles. Thus, MRLs in the most recent toxicological profiles supersede previously published levels. For additional information regarding MRLs, please contact the Division of Toxicology, Agency for Toxic Substances and Disease Registry, 1600 Clifton Road, Mailstop E-29, Atlanta, Georgia 30333.
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APPENDIX A
MINIMAL RISK LEVEL (MRL) WORKSHEET
Chemical name: Uranium (Soluble Forms) CAS number: Multiple Date: July 2001 Profile status: Final Route: [x] Inhalation [ ] Oral Duration: [ ] Acute [x] Intermediate [ ] Chronic Key to figure: 72 Species: Dog
MRL: 4×10-4 [ ] mg/kg/day [ ] ppm [x] mg/m3
Reference: Rothstein A 1949a. Uranyl Fluoride. In: Voegtlin C, Hodge HC, eds. Pharmacology and toxicology of uranium compounds. National Nuclear Energy Series: Manhattan Project Technical Section, Division VI, Vol 1. New York, NY: McGraw-Hill. pp 548-560.
Experimental design: (human study details or strain, number of animals per exposure/control group, sex, dose administration details):
Dogs (2–6 per group; sex and strain not specified) were exposed to 0.19, 2.8, or 12.2 mg/m3 of uranyl fluoride dust (0.15, 2.2, or 9.2 mg U/m3 ) for 6 hours/day, 6 days/week for 5 weeks. (Doses were analytically determined, not estimated.) Dogs were bodily exposed to the dust. The activity median aerodynamic diameter (AMAD) for the particles is assumed to be 1.5–2.1 μm; average 1.8 μm (see Pozzani 1949). Clinical signs of toxicity, mortality, body weight changes, hematology, blood and urine chemistries were monitored. At the termination of the study, the animals were sacrificed and selected organs were histopathologically examined and uranium levels determined.
Effects noted in study and corresponding doses: Severe toxicity was observed at the highest concentration
(9.2 mg U/m3 ) leading to death. The 2 animals in this group showed signs of anorexia, rhinitis, and polydipsia. Later, these animals vomited blood, had severe weight loss and muscle weakness, and exhibited lassitude prior to death. Histopathological examination of the kidney revealed “severe” tubular lesions. Dogs exposed to 0.15 or 2.2 mg U/m3 (6 per group) had no clinical signs of toxicity or significant weight changes. Clinical chemistry results included increased blood nonprotein nitrogen (NPN) with the maximum value over 200 mg/L in the 2 dogs exposed to 9.2 mg U/m3. At 0.15 mg U/m3, blood NPN and urinary amino acid nitrogen were normal in 3 dogs, while 1 of the 3 had increased urinary protein (not all tests were run on all dogs). Histopathological examination of the kidneys revealed “moderate” damage at 2.2 mg U/m3and “slight” changes in 50% of the dogs at 0.15 mg U/m3.
Dose endpoint used for MRL derivation:
[ ] NOAEL [x] LOAEL
0.15 mg/m3; minimal microscopic lesions in the renal tubules in half the dogs examined. Proteinuria observed at 2.2 mg/m3, severe renal damage at 9.2 mg/m3.
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APPENDIX A
Uncertainty factors used in MRL derivation:
[ ] 1 [x] 3 [ ] 10 (for use of a LOAEL)[ ] 1 [x] 3 [ ] 10 (for extrapolation from animals to humans)[ ] 1 [ ] 3 [x] 10 (for human variability)
Was a conversion factor used from ppm in food or water to a mg/body weight dose?If so, explain: Not applicable.
If an inhalation study in animals, list conversion factors used in determining human equivalent dose: See calculations.
Was a conversion used from intermittent to continuous exposure? Yes. See calculations.
Other additional studies or pertinent information that lend support to this MRL:The United Nations Scientific Committee on the Effects of Atomic Radiation (UNSCEAR) has stated thatlimits for natural (and depleted) uranium in drinking water (the most important source of human exposure)should be based on the chemical toxicity rather than on a hypothetical radiological toxicity in skeletaltissues, which has not been observed in either man or animals (Wrenn et al. 1985).
Uranium is a nephrotoxin, exerting its toxic effect by chemical action mostly in the proximal tubule in humans and animals.
Numerous intermediate-duration uranium exposure studies in animals show that the most sensitive effect is renal toxicity (Dygert 1949a, 1949b, 1949c; Pozzani 1949; Roberts 1949; Rothermel 1949; Rothstein 1949a, 1949c, 1949d; Spiegl 1949; Stokinger et al. 1953). Dogs and rabbits appear to be the most sensitive species while rats are less sensitive. Susceptibility also depends on the chemical form of the uranium, the more water-soluble compounds being more toxic than the insoluble compounds. Nephrotoxic effects found in these animals range from minimal tubular lesions without functional effects to proteinuria, acute tubular necrosis, and renal failure.
Calculations
Since inhalation MRL's are derived for continuous exposure, the animal LOAEL derived from an intermittent exposure must be adjusted to continuous exposure:
For an exposure of 6 hours a day, 6 days a week,
(0.15 mg/m3) * (6/24) * (6/7) = 0.032 mg/m3 adjusted to continuous exposure.
URANIUM A-5

APPENDIX A
The adjusted animal NOAEL(ADJ) must be converted to Human Equivalent Concentration (LOAELHEC) before applying uncertainty factors (UFs) adjustments (EPA 1994) for the derivation of the inhalation MRL:
NOAEL[HEC] ' NOAEL[ADJ] x RDDR
where:
NOAEL[ADJ] = duration adjusted laboratory animal NOAEL (in mg/m3).
NOAEL[HEC] = human equivalent concentration of adjusted laboratory animal dose (in mg/m3).
RDDR = Regional Deposited Dose Ratio
Since RDDR values are unavailable for dogs (EPA 1994), ATSDR used a default uncertainty factor of 3 for extrapolating from animals to humans as it incorporates the differences in physiology between dogs and humans. A default factor of 3 was used rather than the standard factor of 10 because of similarities in renal physiology between the two species, i.e., both acidify the urine by active transport of bicarbonate. Additional uncertainty factors of 3 for use of a minimal LOAEL and 10 for human intraspecies variability are used to calculate the intermediate-duration intermediate MRL.
NOAEL(HEC)Intermediate Inhalation MRL ' j UF Adjustments
Therefore,
Intermediate Inhalation MRL ' 0.032 mg/m 3 ' 4 x 10&4 mg/m3. 90
Agency Contact (Chemical Manager): Sam Keith.
URANIUM A-6

APPENDIX A
MINIMAL RISK LEVEL (MRL) WORKSHEET
Chemical name: Uranium (Insoluble Forms) CAS number: Multiple Date: July 2001 Profile status: Final Route: [x] Inhalation [ ] Oral Duration: [ ] Acute [x] Intermediate [ ] Chronic Key to figure: 73 Species: Dog
MRL: 8×10-3 [ ] mg/kg/day [ ] ppm [x] mg/m3
Reference: Rothstein A. 1949b. Uranium Dioxide. In: Voegtlin C, Hodge HC, eds. Pharmacology and toxicology of uranium compounds. National Nuclear Energy Series: Manhattan Project Technical Section, Division VI, Vol 1. New York, NY: McGraw-Hill. pp 614-621.
Experimental design: (human study details or strain, number of animals per exposure/control group, sex, dose administration details):
Dogs (N=6–19; unspecified sex and strain) were exposed to uranium dioxide dust at concentrations of
1.1 mg U/m3, 8.2 mg U/m3, or 9.2 mg U/m3 for 5 weeks, 6 days/weeks, 6 hours a day. (Doses were analytically determined, not estimated.) Studies conducted at 8.2 mg U/m3 were conducted in head exposure units. Studies conducted at the other concentrations were performed in full exposure units. The activity median aerodynamic diameter (AMAD) for the particles is assumed to be 1.5–2.1 μm; average 1.8 μm (see Pozzani 1949). Mortality, body weight changes, standard hematology (except in the 8.2 mg U/m3 group), blood and urine chemistries, pathology, and uranium distribution in tissues were measured.
Effects noted in study and corresponding doses: No dogs died from exposure to uranium dioxide dust. Additionally, no significant weight changes, or biochemical changes in blood or urine were seen at any concentration. No hematological changes were attributable to uranium dioxide dust. Histopathological changes in the kidney were not observed in any group except for “very slight” renal tubular degeneration in 2 of 6 dogs at 8.2 mg U/m3.
Dose endpoint used for MRL derivation:
[x] NOAEL [] LOAEL
1.1 mg/m3; (LOAEL for minimal microscopic lesions in the renal tubules observed at 8.2 mg/m3 in 2 of 6 dogs examined).
URANIUM A-7

APPENDIX A
Uncertainty factors used in MRL derivation:
[ ] 1 [ ] 3 [ ] 10 (for use of a LOAEL)[ ] 1 [x] 3 [ ] 10 (for extrapolation from animals to humans)[ ] 1 [ ] 3 [x] 10 (for human variability)
Was a conversion factor used from ppm in food or water to a mg/body weight dose?If so, explain: Not applicable.
If an inhalation study in animals, list conversion factors used in determining human equivalent dose: See calculations.
Was a conversion used from intermittent to continuous exposure? Yes. See calculations.
Other additional studies or pertinent information that lend support to this MRL:The United Nations Scientific Committee on the Effects of Atomic Radiation (UNSCEAR) has stated thatlimits for natural (and depleted) uranium in drinking water (the most important source of human exposure)should be based on the chemical toxicity rather than on a hypothetical radiological toxicity in skeletaltissues, which has not been observed in either man or animals (Wrenn et al. 1985).
Uranium is a nephrotoxin, exerting its toxic effect by chemical action mostly in the proximal tubule in humans and animals.
Numerous intermediate-duration uranium exposure studies in animals show that the most sensitive effect is renal toxicity (Dygert 1949a, 1949b, 1949c; Pozzani 1949; Roberts 1949; Rothermel 1949; Rothstein 1949a, 1949c, 1949d; Spiegl 1949; Stokinger et al. 1953). Dogs and rabbits appear to be the most sensitive species while rats are less sensitive. Susceptibility also depends on the chemical form of the uranium, the more water-soluble compounds being more toxic than the insoluble compounds. Nephrotoxic effects found in these animals range from minimal tubular lesions without functional effects to proteinuria, acute tubular necrosis, and renal failure.
Calculations
Since inhalation MRL's are derived for continuous exposure, the animal NOAEL derived from an intermittent exposure must be adjusted to continuous exposure:
For an exposure of 6 hours a day, 6 days a week,
(1.1 mg/m3) * (6/24) * (6/7) = 0.2357 mg/m3 adjusted to continuous exposure.
URANIUM A-8

APPENDIX A
The adjusted animal NOAEL(ADJ) must be converted to Human Equivalent Concentration (LOAELHEC) before applying uncertainty factors (UFs) adjustments (EPA 1994) for the derivation of the inhalation MRL:
NOAEL[HEC] ' NOAEL[ADJ] x RDDR
where:
NOAEL[ADJ] = duration adjusted laboratory animal NOAEL (in mg/m3).
NOAEL[HEC] = human equivalent concentration of adjusted laboratory animal dose (in mg/m3).
RDDR = Regional Deposited Dose Ratio
Since RDDR values are unavailable for dogs (EPA 1994), ATSDR used a default uncertainty factor of 3 for extrapolating from animals to humans as it incorporates the differences in physiology between dogs and humans. A default factor of 3 was used rather than the standard factor of 10 because of similarities in renal physiology between the two species, i.e., both acidify the urine by active transport of bicarbonate. An additional uncertainty factor of 10 for human intraspecies variability is used to calculate the intermediate-duration inhalation MRL:
NOAEL(HEC)Intermediate Inhalation MRL ' j UF Adjustments
Therefore,
Intermediate Inhalation MRL ' 0.2357 mg/m 3 ' 8 x 10&3 mg/m3. 30
Agency Contact (Chemical Manager): Sam Keith.
URANIUM A-9

APPENDIX A
MINIMAL RISK LEVEL (MRL) WORKSHEET
Chemical name: Uranium (Soluble forms) CAS number: Multiple Date: July 2001 Profile status: Final Route: [x] Inhalation [ ] Oral Duration: [ ] Acute [ ]Intermediate [x] Chronic Key to figure: 112 Species: Dog
MRL: 3×10-4 [ ] mg/kg/day [ ] ppm [x] mg/m3
Reference: Stokinger et al. 1953. Uranium Tetrachloride: Toxicity following inhalation for 1 and 2 years. In: Voegtlin C, Hodge HC, eds. Pharmacology and toxicology of uranium compounds. National Nuclear Energy Series: Manhattan Project Technical Section, Division VI, Vol 1. New York, NY: McGraw-Hill. pp 1522-1553.
Experimental design: (human study details or strain, number of animals per exposure/control group, sex, dose administration details):
Dogs of both sexes (11-12 M, 9-10 F) were exposed to uranium tetrachloride in inhalation chambers for 6 hours a day, M-F and 3 hours on Saturday (5.5 days a week) for 1 year at concentrations of 0, 0.05, and
0.20 mg U/m3. (Doses were analytically determined, not estimated.) The activity median aerodynamic diameter (AMAD) of the aerosols was 1–2 μm. The animals were monitored for body weight alterations, clinical signs of toxicity, and biochemical alterations in the blood and urine. At the termination of the study, the animals were sacrificed and selected organs were histopathologically examined.
Effects noted in study and corresponding doses: All dogs survived the 1-year exposure period. No significant changes were observed in blood non-protein nitrogen, hematology, histopathology of liver, body weight, or urinary proteins. “Very slight” renal damage as reported in animals exposed to 0.20 mg U/m3. Histological and biochemical examinations revealed a NOAEL level of 0.05 mg U/m3 and minimal microscopic lesions in the renal tubules in the 0.20 mg U/m3 dose level dogs. No significant weight loss was observed in the dogs.
Dose endpoint used for MRL derivation:
[x] NOAEL [] LOAEL
0.05 mg/m3; (minimal microscopic lesions in the renal tubules in dogs of both sexes at a LOAEL of
0.20 mg U/m3).
Uncertainty factors used in MRL derivation:
[ ] 1 [ ] 3 [ ] 10 (for use of a LOAEL)[ ] 1 [x] 3 [ ] 10 (for extrapolation from animals to humans)[ ] 1 [ ] 3 [x] 10 (for human variability)
URANIUM A-10

APPENDIX A
Modification factors used in MRL derivation:
[ ] 1 [ ] 3 [ ]10
Was a conversion factor used from ppm in food or water to a mg/body weight dose? If so, explain: Not applicable.
If an inhalation study in animals, list conversion factors used in determining human equivalent dose: See calculations.
Was a conversion used from intermittent to continuous exposure? Yes. See calculations.
Other additional studies or pertinent information that lend support to this MRL:The United Nations Scientific Committee on the Effects of Atomic Radiation (UNSCEAR) stated that limitsfor natural (and depleted) uranium in drinking water (the most important source of human exposure) shouldbe based on the chemical toxicity rather than on a hypothetical radiological toxicity in skeletal tissues,which has not been observed in either man or animals (Wrenn et al. 1985).
Uranium has been identified as a weak metal nephrotoxin, exerting its toxic effect by chemical action mostly in the proximal tubule in humans and animals. A study on the kidney functions of uranium mill workers chronically exposed to insoluble uranium (uranium oxide) showed renal tubular dysfunction (as manifested by mild proteinuria, aminoaciduria, and a dose-related clearance of β-2-microglobulin) relative to that of creatinine and the length of time that the uranium workers had spent in the yellowcake (uranium oxides) drying and packaging area. Serum β-2-microglobulin was also elevated in the serum of 22 of the 23 workers tested (Saccomanno et al. 1982; Thun 1985). However, a histopathological autopsy study of individuals, who had been occupationally exposed uranium workers and then spent many years in retirement, found that the damage potentially caused by internalized uranium during their years of occupational exposure had been repaired and was not detectable at death (Russell et al. 1996)
In animal studies, chronic-duration studies with rat and dogs given inhalation uranium (uranium tetrachloride, uranium tetrafluoride, uranyl nitrate hexahydrate, uranium dioxide) doses as low as
0.05 mg U/m3 and as high as 10 mg U/m3, administered for 1–5 years suffered nephrotoxicity. Nephrotoxic effects found in these animals ranged from proteinuria and increased bromosulfalein retention (for low doses) (Stokinger et al. 1953) to acute tubular necrosis (for high doses) (Leach et al. 1970). No treatment-related renal effects were seen when Rhesus monkeys and dogs were exposed to uranium dioxide by inhalation at doses as high as 5.1 mg U/m3 for 1–5 years (Leach et al. 1973). Guinea pigs, mice, rat, cats, rabbits, and dogs (Dygert 1949a, 1949b, 1949c; Pozzani 1949; Roberts 1949; Rothermel 1949; Rothstein 1949a, 1949c, 1949d; Spiegl 1949; Stokinger et al. 1953), or guinea pigs, rabbits, and rats (Leach et al. 1984; Morrow et al. 1982; Roberts 1949; Stokinger et al. 1953) exposed to these uranium compounds in intermediate- or acute-duration exposure and rats and dogs in chronic-duration studies (Leach et al. 1970; Stokinger et al. 1953) suffered similar kidney damage.
Calculations
Since inhalation MRL's are derived for continuous exposure, the animal NOAEL derived from an intermittent exposure must be adjusted to continuous exposure:
URANIUM A-11

APPENDIX A
NOAELADJ ' 0.05 mg uranium/m 3 x (6) hours x (5.5) days ' 0.01 mg uranium/m3. 247
The adjusted animal NOAEL(ADJ) must be converted to Human Equivalent Concentration (LOAELHEC) before applying uncertainty factors (UFs) adjustments (EPA 1994) for the derivation of the inhalation MRL:
NOAEL[HEC] ' NOAEL[ADJ] x RDDR
where:
NOAEL[ADJ] = duration adjusted laboratory animal NOAEL (in mg/m3).
NOAEL[HEC] = human equivalent concentration of adjusted laboratory animal dose (in mg/m3).
RDDR = Regional Deposited Dose Ratio
Since RDDR values are unavailable for dogs (EPA 1994), ATSDR used a default uncertainty factor of 3 for extrapolating from animals to humans as it incorporates the differences in physiology between dogs and humans. A default factor of 3 was used rather than the standard factor of 10 because of similarities in renal physiology between the two species, i.e., both acidify the urine by active transport of bicarbonate. An additional uncertainty factor of 10 for human intraspecies variability is used to calculate the chronic
inhalation MRL:
Chronic Inhalation MRL ' NOAEL(HEC)j UF Adjustments
Therefore,
Chronic Inhalation MRL ' 0.01 mg/m3 30 ' 3 x10&4 mg/m3 .

Agency Contact (Chemical Manager): Sam Keith.
URANIUM A-12

APPENDIX A
MINIMAL RISK LEVEL (MRL) WORKSHEET
Chemical name: Uranium CAS number: Multiple Date: July 2001 Profile status: Final Route: [ ] Inhalation [x] Oral Duration: [ ] Acute [x] Intermediate [ ] Chronic Key to figure: 57 Species: Rabbit
MRL: 2×10-3 [x] mg/kg/day [ ] ppm [ ] mg/m3
Reference: Gilman AP, Villeneuve DC, Secours VE, et al. 1998b. Uranyl nitrate – 91-day toxicity studies in the New Zealand white rabbit Toxicol Sci. 41(1):129-137, Jan 1998.
Experimental design (human study details or strain, number of animals per exposure/control group, sex, dose administration details):
Groups of New Zealand rabbits (10/sex/dose, males 3,200 g, females 3,100 g) were exposed to uranium as uranyl nitrate in drinking water. The males were exposed to 0.96, 4.8, 24, 120, or 600 mg/L for 91 days, while females were exposed only to 4.8, 24 or 600 mg/L. A control group of 10 animals for each sex received tap water without uranyl nitrate (<0.001 mg U/L). All animals were fed chow containing greater than > greater than or equal to = equal to < less than 20 years. The mechanism by which cancer is induced in living cells is complex and is a topic of intense study. Exposure to ionizing radiation can produce cancer at any site within the body; however, some sites appear to be more common than others, such as the breast, lung, stomach, and thyroid.
DNA is a major target molecule during exposure to ionizing radiation. Other macromolecules, such as lipids and proteins, are also at risk of damage when exposed to ionizing radiation. The genotoxicity of ionizing radiation is an area of intense study, as damage to the DNA is ultimately responsible for many of the adverse toxicological effects ascribed to ionizing radiation, including cancer. Damage to genetic material is basic to developmental or teratogenic effects, as well. However, for effects other than cancer, there is little evidence of human effects at low levels of exposure.
D.5 UNITS IN RADIATION PROTECTION AND REGULATION
D.5.1 Dose Equivalent and Dose Equivalent Rate.
Dose equivalent or rem is a special radiation protection quantity that is used, for administrative and radiation safety purposes only, to express the absorbed dose in a manner which considers the difference in biological effectiveness of various kinds of ionizing radiation. The ICRU has defined the dose equivalent, H, as the product of the absorbed dose, D, and the quality factor, Q, at the point of interest in biological tissue. This relationship is expressed as H = D x Q. The dose equivalent concept is applicable only to doses that are not great enough to produce biomedical effects.
URANIUM D-10

APPENDIX D
The quality factor is a dimensionless quantity that depends in part on the stopping power for charged particles, and it accounts for the differences in biological effectiveness found among the types of radiation. Originally relative biological effectiveness (RBE) was used rather than Q to define the quantity, rem, which was of use in risk assessment. The generally accepted values for quality factors for various radiation types are provided in Table D-3. The dose equivalent rate is the time rate of change of the dose equivalent to organs and tissues and is expressed as rem/unit time or sievert/unit time.
Table D-3. Quality Factors (Q) and Absorbed Dose Equivalencies
Type of radiation Quality factor (Q) Absorbed dose equal to a unit dose equivalent*
X, gamma, or beta radiation 1 1
Alpha particles, multiple-charged particles, fission fragments and heavy particles of unknown charge 20 0.05
Neutrons of unknown energy 10 0.1
High-energy protons 10 0.1

* Absorbed dose in rad equal to 1 rem or the absorbed dose in gray equal to 1 sievert.
Source: USNRC. 1999. Standards for the protection against radiation, table 1004(b).1. 10 CFR 20.1004. U.S. Nuclear Regulatory Commission, Washington, D.C.
D.5.2 Relative Biological Effectiveness.
RBE is used to denote the experimentally determined ratio of the absorbed dose from one radiation type to the absorbed dose of a reference radiation required to produce an identical biologic effect under the same conditions. Gamma rays from cobalt-60 and 200–250 keV x-rays have been used as reference standards. The term RBE has been widely used in experimental radiobiology, and the term quality factor used in calculations of dose equivalents for radiation safety purposes (ICRP 1977; NCRP 1971; UNSCEAR 1982). RBE applies only to a specific biological end point, in a specific exposure, under specific conditions to a specific species. There are no generally accepted values of RBE.
D.5.3 Effective Dose Equivalent and Effective Dose Equivalent Rate.
The absorbed dose is usually defined as the mean absorbed dose within an organ or tissue. This representsa simplification of the actual problem. Normally when an individual ingests or inhales a radionuclide or isexposed to external radiation that enters the body (gamma), the dose is not uniform throughout the wholebody. The simplifying assumption is that the detriment will be the same whether the body is uniformly ornon-uniformly irradiated. In an attempt to compare detriment from absorbed dose of a limited portion ofthe body with the detriment from total body dose, the ICRP (1977) has derived a concept of effective doseequivalent. The effective dose equivalent, HE, is
HE = (the sum of) Wt Ht
URANIUM D-11

APPENDIX D
where Ht is the dose equivalent in the tissue, Wt is the weighting factor, which represents the estimatedproportion of the stochastic risk resulting from tissue, T, to the stochastic risk when the whole body isuniformly irradiated for occupational exposures under certain conditions (ICRP 1977). Weighting factorsfor selected tissues are listed in Table D-4.
The ICRU (1980), ICRP (1984), and NCRP (1985) now recommend that the rad, roentgen, curie, and rem be replaced by the SI units: gray (Gy), Coulomb per kilogram (C/kg), Becquerel (Bq), and sievert (Sv), respectively. The relationship between the customary units and the international system of units (SI) for radiological quantities is shown in Table D-5.
Table D-4. Weighting Factors for Calculating Effective Dose Equivalent for Selected Tissues
Weighting factor
Tissue ICRP60 NCRP115/ ICRP60 NRC
Bladder 0.040 0.05 –
Bone marrow 0.143 0.12 0.12
Bone surface 0.009 0.01 0.03
Breast 0.050 0.05 0.15
Colon 0.141 0.12 –
Liver 0.022 0.05 –
Lung 0.111 0.12 0.12
Esophagus 0.034 0.05 –
Ovary 0.020 0.05 –
Skin 0.006 0.01 –
Stomach 0.139 0.12 –
Thyroid 0.021 0.05 0.03
Gonads 0.183 0.20 0.25
subtotal 0.969 1 0.70
Remainder 0.031 0.05 0.30

ICRP60 = International Commission on Radiological Protection, 1990 Recommendations of the ICRP;NCRP115 = National Council on Radiation Protection and Measurements. 1993. Risk Estimates for RadiationProtection, Report 115. Bethesda, Maryland; NRC = Nuclear Regulatory Commission.NRC = Nuclear Regulatory Commission, Title 10, Code of Federal Regulations, Part 20
URANIUM D-1 2APPENDIX D

Table D-5. Comparison of Common and SI Units for Radiation Quantities
Quantity Customary units Definition SI units Definition
Activity (A) curie (Ci) 3.7×1010 transformations s-1 becquerel (Bq) s-1
Absorbed dose (D) rad (rad) 10-2 Jkg-1 gray (Gy) Jkg-1
Absorbed dose rate (Š) rad per second (rad s-1) 10-2 Jkg-1s-1 gray per second (Gy s-1) Jkg-1 s-1
Dose equivalent (H) rem (rem) 10-2 Jkg-1 sievert (Sv) Jkg-1
Dose equivalent rate ( ) rem per second (rem s-1) 10-2 Jkg-1s-1 sievert per second (Sv s-1) Jkg-1 s-1
Linear energy transfer (LET) kiloelectron volts per micrometer (keV μm-1) 1.602×10-10 Jm-1 kiloelectron volts per micrometer (keV μm-1) 1.602×10-10 Jm-1

Jkg-1 = Joules per kilogram; Jkg-1s-1 = Joules per kilogram per second; Jm-1 = Joules per meter; s-1 = per second
URANIUM D-13

APPENDIX D
REFERENCES FOR APPENDIX D
ATSDR. 1990a. Toxicological profile for thorium. U.S. Department of Health and Human Services. Public Health Service. Agency for Toxic Substances and Disease Registry. Atlanta, GA.
ATSDR. 1990b. Toxicological profile for radium. U.S. Department of Health and Human Services. Public Health Service. Agency for Toxic Substances and Disease Registry. Atlanta, GA.
ATSDR. 1990c. Toxicological profile for radon. U.S. Department of Health and Human Services. Public Health Service. Agency for Toxic Substances and Disease Registry. Atlanta, GA.
ATSDR. 1999. Toxicological profile for uranium. U.S. Department of Health and Human Services. Public Health Service. Agency for Toxic Substances and Disease Registry. Atlanta, GA.
BEIR III. 1980. The effects on populations of exposure to low levels of ionizing radiation. Committee on the Biological Effects of Ionizing Radiations, National Research Council. Washington, DC: National Academy Press.
BEIR IV. 1988. Health risks of radon and other internally deposited alpha emitters. Committee on the Biological Effects of Ionizing Radiations, National Research Council. Washington, DC: National Academy Press.
BEIR V. 1988. Health effects of exposure to low levels of ionizing radiation. Committee on the Biological Effects of Ionizing Radiations, National Research Council. Washington, DC: National Academy Press.
Brodsky A. 1996. Review of radiation risks and uranium toxicity with application to decisions associated with decommissioning clean-up criteria. Hebron,Connecticut: RSA Publications.
Cember H. 1996. Introduction to health physics. New York., NY: McGraw Hill.
Early P, Razzak M, Sodee D. 1979. Nuclear medicine technology. 2nd ed. St. Louis: C.V. Mosby Company.
Eichholz G. 1982. Environmental aspects of nuclear power. Ann Arbor, MI: Ann Arbor Science.
Hendee W. 1973. Radioactive isotopes in biological research. New York, NY: John Wiley and Sons.
Hobbs C, McClellan R. 1986. Radiation and radioactive materials. In: Doull J, et al., eds. Casarett and Doull’s Toxicology. 3rd ed. New York, NY: Macmillan Publishing Co., Inc., 497-530.
ICRP. 1977. International Commission on Radiological Protection. Recommendations of the International Commission on Radiological Protection. ICRP Publication 26. Vol 1. No. 3. Oxford: Pergamon Press.
ICRP. 1979. International Commission on Radiological Protection. Limits for intakes of radionuclides by workers. ICRP Publication 20. Vol. 3. No. 1-4. Oxford: Pergamon Press.
URANIUM D-14

APPENDIX D
ICRP. 1979. Limits for Intakes of Radionuclides by Workers. Publication 30. International Commission
on Radiological Protection. Pergamon Press.ICRP. 1984. International Commission on Radiological Protection. A compilation of the major conceptsand quantities in use by ICRP. ICRP Publication 42. Oxford: Pergamon Press.
ICRP. 1990. International Commission on Radiological Protection 1990 Recommendations of the ICRP
ICRU. 1980. International Commission on Radiation Units and Measurements. ICRU Report No. 33. Washington, DC.James A. 1987. A reconsideration of cells at risk and other key factors in radon daughter dosimetry. In:
Hopke P, ed. Radon and its decay products: Occurrence, properties and health effects. ACS Symposium
Series 331. Washington, DC: American Chemical Society, 400-418.James A, Roy M. 1987. Dosimetric lung models. In: Gerber G, et al., ed. Age-related factors inradionuclide metabolism and dosimetry. Boston: Martinus Nijhoff Publishers, 95-108.
Kondo S. 1993. Health effects of low-level radiation. Kinki University Press, Osaka, Japan (available
from Medical Physics Publishing, Madison, Wisconsin).Kato H, Schull W. 1982. Studies of the mortality of A-bomb survivors. Report 7 Part 8, Cancer mortalityamong atomic bomb survivors, 1950-78. Radiat Res 90;395-432.
Mettler F, Moseley R. 1985. Medical effects of ionizing radiation. New York: Grune and Stratton.
NCRP. 1971. Basic radiation protection criteria. National Council on Radiation Protection andMeasurements. Report No. 39. Washington, DC.NCRP. 1985. A handbook of radioactivity measurements procedures. 2nd ed. National Council on
Radiation Protection and Measurements. Report No. 58. Bethesda, MD:
NCRP. 1993. Risk estimates for radiation protection. National Council on Radiation Protection andMeasurements. Report 115. Bethesda, MarylandOtake M, Schull W. 1984. Mental retardation in children exposed in utero to the atomic bombs: A
reassessment. Technical Report RERF TR 1-83, Radiation Effects Research Foundation, Japan.Rubin P, Casarett G. 1968. Clinical radiation pathology. Philadelphia: W.B. Sanders Company, 33.UNSCEAR. 1977. United Nations Scientific Committee on the Effects of Atomic Radiation. Sources and
effects of ionizing radiation. New York: United Nations.
UNSCEAR. 1982. United Nations Scientific Committee on the Effects of Atomic Radiation. Ionizingradiation: Sources and biological effects. New York: United Nations.UNSCEAR. 1986. United Nations Scientific Committee on the Effects of Atomic Radiation. Genetic and
somatic effects of ionizing radiation. New York: United Nations.UNSCEAR. 1988. United Nations Scientific Committee on the Effects of Atomic Radiation. Sources,effects and risks of ionization radiation. New York: United Nations.

Uranium – DU or Otherwise- safe or unsafe

December 23, 2010

And uranium decays to much more radioactive progeny. The older the manufactured uranium is, the more radium, polonium etc will be present within the mass. And of course radon is vented from the mass.

Anyhow, regardless of what Bobby Scott, Lovelace Institute, Los Alamos and Flinders University would have you believe about the supposed benefits of low dose radiation, here is the current US Toxicological Profile for Uranium published by the US Agency for Toxic Substances and Disease Register.

Each part is a downloadable pdf.

I might reproduce some of it here.

Happy Christmas or whatever the time of year is to you.

Uranium isnt safe. I live in the uranium capital of the Southern Hemisphere – South Australia. The Powers would have me believe the mining of it is safe and that the use of it is safe.

The facts show that the state I live in is a becoming a bigger and bigger hole in the ground for the sake of the military industrial complex.

I guess that makes me a “feral” in the eyes of the state treasurer, Kevin Foley.

Maybe he should wander round the pubs and clubs late at night less and think in the light a lot more. Though, I must admit, I’d love to chain to a chair at my place, get him drunk and do a “Clockwork Orange” on him in front of my DVD player.

That’s after all, wot he does to us.

And we pay his wages. Stupid is as stupid does I guess. Im no genius.

I do know this much. We employ Foley. The only name he is reasonably able to call us, not matter what our individual views, is “Voter”.

He is the feral. In the real sense of the world. A life form of exotic origins damaging to the local ecosystem.

Apple iMac 2006-2007 Monitor Death – Solution

December 23, 2010

Change of pace. Couldnt write the blog without my Macs.

Yesterday the lcd screen of my late 2006 intel iMac died. Tried setting up an external monitor using the mini dvi to vga plug that is available. Trouble is OSX defaults to split screen, not mirror mode. So though the external monitor worked, it didnt have the icons on it, couldnt access System Preferences to change to mirror mode.

The solution was to remove the front bezel, and unplug the two cables that go from the LCD panel to the imac main board. I put it all back together and reconnected the external monitor. It worked no worries. OSX only sees the external monitor and boots up with that as the main monitor. Problem solved. So now I have a two piece iMac. Big Deal. The main thing is it works.

If this has happened to you dont take the iMac apart unless you know what you are doing. The take apart manual is here:

Mac Manuals

And follow all the needed procedures and precautions.

This method worked for me, I share it here as an aid. If you do it incorrectly you may damage your machine and that is your doing, not mine. Best of luck.

The Continuing Cost of Chernobyl

December 16, 2010

“The Australian” newspaper published the following on 15 Dec 2010:

HISTORY OF CHERNOBYL
– On April 26, 1986, a power surge during an unauthorised systems test at Chernobyl’s reactor number 4 sparked explosions and ruptured a reactor vessel. A plume of radioactive fallout drifted over large parts of the western Soviet Union and much of Europe

– The Chernobyl disaster is considered the world’s worst nuclear power plant accident and is the only level-seven event on the International Nuclear Event Scale

– By December 2000, 350,400 people were evacuated and resettled after the disaster, which released the destructive potential of at least 100 atomic bombs. The UN says 56 people died soon after the accident and that thousands are likely to die from radiation contamination

– About 600 power station employees as well as firefighters and soldiers, who were sent to clean up the site, suffered (and still suffer) from conditions including lung cancer and leukemia, cardiovascular diseases and inflammation of the digestive tract

– Between 1986 and 1990, Ukraine recorded an increase in miscarriages, premature births and stillbirths, as well as three times the normal rate of deformities and developmental abnormalities in newborns

– Belarus: Almost a quarter of Belarus remains contaminated.

– The damage for the republic is estimated at $US235 billion, about 60 times the annual national budget

– Ukraine: About 6 per cent of the country remains contaminated. Ukraine still spends about 7 per cent of the national budget dealing with the consequences

– Russia: About 1.5 per cent remains contaminated

This is completely at loggerheads with the crap Bobby Scott (Lovelace Institute, contractor to the US DOE) has published. His apparent main beef about the disaster is the number of voluntary abortions performed in Europe in the aftermath. An ascertain which is demolished by qualified articles to the contary. Chernobyl has and had a huge human and economic toll.

As usual with nuclear sites, there are moves to turn it into a tourist destination.
One wonders when Hanford will be thus similarly treated. Perhaps visitors can feast on traditional Native American tucker such as radioactive rabbits.

The Australian continues:Chernobyl to become tourism hotspot
* James Marson* From: The Australian* December 15, 2010 8:28AM

UKRAINE’S government wants to turn Chernobyl, the site of the world’s worst nuclear accident, into a tourism hotspot.

Ukraine’s Emergency Situations Ministry said it was working on a plan to open the area around the defunct plant – where a reactor exploded on April 26, 1986, spreading radiation across the then Soviet states of Ukraine, Belarus and Russia – to visitors starting next month.

The ministry said radiation levels in certain parts of the so-called exclusion zone, which stretches 30km around the exploded reactor, were now returning to normal levels, The Australian reports.

Visitors will be able to take in views of the nuclear plant, as well as towns and villages that were abandoned in the disaster’s aftermath.

Official tour operators would have to meet strict criteria to be allowed to operate, said Yulia Yurshova, a spokeswoman for the Emergency Situations Ministry, as straying from the route can be dangerous because of the threat of collapsing buildings and varying radiation levels.

“The Chernobyl zone isn’t as scary as the whole world thinks,” Ms Yurshova said.
“We want to work with big tour operators and attract Western tourists, from whom there’s great demand.”
About 2500 people still maintain the plant.

Tours to Chernobyl and the sealed area around the plant – many of which are run illegally, according to Ms Yurshova – attract about 6000 visitors a year and cost about $150.70 for a day trip.
Ms Yurshova said official tours would begin next month.
UN Development Program leader Helen Clark supported the plan.

“There is an opportunity to tell a story here and of course the process of telling a story, even a sad story, is something that is positive in economic terms and positive in conveying very important messages,” she said.

Work on a new sarcophagus to seal the reactor is expected to be completed by 2015, the ministry also announced.

The shield, made of metal and concrete, will cost $US1.2 billion and will be financed by Ukraine and international donors.

http://www.theaustralian.com.au/travel/world/chernobyl-to-become-tourism-hotspot/story-fn302659-1225971300611

LD 50 – variations in responses to radiation exposures

December 12, 2010

The following is taken from Chapter 3 of “Atomic Radiation and Life”, by Peter Alexander,
Pelican Books, London, 1957.

It will take me some time to scan and upload all of the text and diagrams of this chapter. Bear with me and the basis of variation in responses to radiation between species, strains and individuals will be explained.

The variable responses produced by variations in metabolism and other variables will also be explained.

These variations between individuals are additionally affected according to the circumstances involving other variables at the time of exposure. These facts show that simple one response per species – including Humans- is inadequate. This is indeed the situation witnessed in the immediate aftermath
Of Hiroshima. Hiroshima doctors noted that those people at rest at the time of exposure were more likely to survive radiation sickness than those who were experting themselves.

Further, the Hiroshima doctors noted those who suffered thermal burns of a non lethal nature were also more likely to survive radiation sickness than those exposed at the same distance from the bomb blast than those who did not suffer similar thermal burns. (Hersey).

The simplistic study of the responses of special mice in response to external x rays –
Will be seen to be inadequate to explain the full range of human responses to exposures to ionizing radiation of lesser or greater absorbed doses. Both internal and external. The DOE Sykes study -upon which Scott basis his statements regarding the supposed health benefits of not cleaning up contaminated sites etc – must receive critical study by scientists not in the pay of DOE and who do
Not have conflicts of interest determined by their paid roles.

CHAPTER 3
RADIATION SICKNESS

Mammals are exceptionally sensitive to radiation. Table III
on p. 74 shows that there are considerable variations in the
radio-sensitivity of different mammals. None of these figures
are absolute as they depend on the strain of the given
animal: there are mice which have been selectively bred
for high or low radiation resistance, and by successive
brother-sister matings so-called pure strains are obtained
which show more uniform behaviour and have a smaller
f
variation (see Fig. 16). The lethal dose may vary by as
much as 30 per cent between these strains. Other factors also
have a minor influence on the radiation sensitivity; it has
been found that in some cases females are able to sustain a
10 per cent greater dose than males. Body weight appears
to have very little influence; this is surprising, since the
total amount of energy deposited by a given dose of radiation
is proportional to the weight of the animal. More
energy, therefore, will be left in a fat than in a lean animal.

Increase in age substantially lowers resistance to radiation;
the lethal dose for 18-month-old rats was 30 per cent less
than that of young adults (three months) from the same
strain. This is a most important finding since it complements
the observation which is discussed in detail on p. 92
that small doses of radiation produce premature ageing.

As already mentioned, animals in hibernation are remarkably
resistant, and doses of many thousands of roentgen are
necessary, to kill marmots or squirrels while they are dormant.

On warming, the animals will behave as if they had
been irradiated in the non-hibernating state. The bat
appeared to be the one remarkable exception amongst mammals
in that the lethal dose observed was of the order of
15,000 r, i.e. it was twenty to fifty times more resistant than
other mammals. Admittedly it is a hibernating species, but
the experiments were carried out at ordinary temperatures
when the animals should be normal. But these bats did not
eat in captivity and this lowered their metabolic rate to a
state equivalent to hibernation and was responsible for the
apparent radiation resistance. Bats which were eating
succumbed to 700 r.

It is now possible to cool small mammals such as mice
to a temperature only two or three degrees above freezing
point and to return them without harm to ordinary temperature
a few hours later. If they are irradiated at these low
temperatures they are truly radiation-resistant (that is,
even after they have been warmed up) and can tolerate two to
three times the dose which normally would kill them.

The reason for this resistance is that their tissues are deficient in
Oxygen when cold, and this protects against radiation (see p.166).

For obvious reasons no accurate data for man are available,
and a value can only be deduced from the casualties
of the atomic bomb explosions at Hiroshima and Nagasaki.

A fairly good estimate of the radiation intensity at different
distances from the explosion centres can be made, and the
approximate dose received by individuals in different
regions can be calculated. The majority of injuries were due
to blast and the frightful burns produced by the intense heat
given off at the flash of the explosion. From victims who
escaped death from these effects information was obtained
about the effects of gamma radiation. Within 1,ooo metres (five eighths
of a mile) the dose received was more than 1,000 r,
and none who were caught in the open survived for more
than one week. In the zone between 1,000 and 1,250 metres
(dose received about 700 r) those fatally injured died within
two months, but there were a few who appeared to have
recovered. From Table 111 it can be seen that in the order

t

of radiation resistance man ranks between the goat and the
mouse, and is roughly as sensitive as a monkey. In this table
the radiation dose is expressed as the lethal dose required to
kill 50 per cent of the animals in 30 days (abbreviated to LD50/30 days).

The reason why we do not choose the dose
required to kill 1oo per cent is that in any group there is a
natural variation in resistance; some animals succumb more
easily than others and occasionally there are animals which
are very much more hardy. This type of behaviour is shown
in Fig. 16.

If in an experiment one attempts to find the dose at which kills all animals
the value recorded will depend on the chance of having included any
of those that are resistant.

In a series of experiments the values recorded for the 100
per cent lethal dose will vary considerably. The dose
required to kill approximately half the animals can be
determined much more precisely, and this value is then
remarkably constant from experiment to experiment. The
reason for having a definite time limit within which death
occurs will become apparent from the discussion on p. 77.

All the figures given apply when the whole of the animal
has been irradiated. Even if only small parts of the body are
not irradiated, such as the tail of a rat, the lethal dose is
substantially increased. This is brought out clearly in
Table iv, where mice were given 1,025 r, a dose substan-

t
tially greater than that necessary to kill them all; 100 per
cent lethality would probably have been obtained with
700-800 r. Yet after shielding different parts of the animals
there were varying numbers of survivors, showing that the
LD50 had been substantially increased by not exposing a
small part of the animal: protecting the spleen is outstandingly
effective and will be discussed again in another
connexion on p. 175. In the experiments shown in Table iv
the different tissues were shielded by enclosing them in a
sheet of lead through which the X-rays cannot penetrate.

The animal can survive tremendous doses to isolated
parts. It is usual practice to irradiate tumours with many
thousands of roentgen of X-rays without hazarding the
general health of the patient. Yet one-tenth of this dose
given to the whole body would be fatal. Even with localized
irradiation exposure of certain parts is more harmful to the
animal than that of others. Rajewsky working in Germany
used a most ingenious method to detect the critical organs.
He shielded the whole body of the rat with lead except for
a narrow slit, the position of which was progressively
changed. The most sensitive organs were those situated in
the abdomen (the liver, spleen, kidney, and part of the
intestine) but it was not possible to pinpoint any one of
these. To kill the animal it was necessary to give doses of the
order of 8,000 r through this narrow slit even when it
exposed part of the critical area. Shielding of parts of the
body from the radiation not only reduces the lethal effects
but with sub-lethal doses cuts down pathological change
such as induction of leukaemia (see Chapter 5).

THE LETHAL DOSE
The statement can probably be accepted that all exposure
to ionizing radiation is harmful. In view of the general background
of radiation (see p. 122) one must not exaggerate
the importance of small doses. It is probably true that prolonged
exposure to sunlight can lead to skin cancer, and is
the reason for the higher incidence of such tumours among
people who work out of doors, but the remedy is clearly not
a mole-like existence without daylight, any more than the
denial of the enormous benefits of atomic energy is the
answer to the problematical dangers of a small increase in
the radiation background.

We have to distinguish between a dose of radiation given
within a relatively short period and chronic irradiation.
Since the body shows a remarkable capacity for recovery
from radiation it can withstand repeated small doses which
if given over a shorter period or even in one irradiation
would be fatal. Since these recovery processes take many
hours there is no difference between a dose of, for example,
600 r given within one minute at the rate of 600 r/min. or
in one hour at 10 r/rnin. However if the rate is cut down to
1 r/min., the radiation symptoms are very much less, and
the dose required to kill within thirty days will be more
than doubled. If the dose rate is reduced still further, say
to I r/hour, then the animal can tolerate this for very long
periods, and a reduction in the total life-span can be
observed only after receiving in toto a dose approximately
five to six times that necessary to produce death within
thirty days if given at a higher dose rate (see Fig. 17).

f
However, there are long-term effects due to genetic changes
which are independent of dose rate. That is to say they
occur to the same extent whether the dose is given rapidly
or slowly. In practice they are more apparent when the
dose is given slowly, since the symptoms of radiation sickness
are then absent or much less severe. These long-term
effects which are transmitted to the progeny of irradiated
animals will be described in the next two chapters.
Radiation sickness, which forms the subject of this
chapter, can be considered in the same way as poisoning;
the consequences may be fatal, or they may give rise to
illness with certain symptoms which disappear in time.
Where a single dose of X-rays of less than 200 r is given
there is no detectable shortening of the life-span in mice or
rats (the animals on which most of the experiments have
f
Figures
f
been carried out), although symptoms of radiation sickness
appear some time after irradiation (see p. 84). However, the
incidence of tumours many months after the irradiation is
greater than in unirradiated animals, and cancer becomes
prominent as a cause of death. In view of the long latent
period between irradiation and the appearance of malignant
growths (see Chapter 5) their effect on the life expectancy
of experimental animals is difficult to determine.
Though almost all of the pioneer doctors using X-rays died
of cancer, most of them lived for twenty to thirty years after
severe over-exposure to radiations. Their lives were shortened,
but not by very much.
With more than 200 r of X-rays death occurs prematurely,
i.e. the lethal action- of the radiation is recognizable.
As the dose is increased beyond this point the time for 50
per cent of the animals to die falls sharply. The complete
dose mortality curve for a strain of white mice irradiated
with X-rays is shown in Figs. 18 and 19. From the first
graph it can be seen that the ~ ~wi5th 3050 r is 100 days.
This is probably less than half the life expectancy of unirradiated
mice, the exact value for which varies somewhat
from laboratory to laboratory. After this the mortality rate
increases rapidly and 500 r gives L D ~ O in thirty days.
Increasing the dose beyond this reduces the time for death
relatively more slowly, and at 1,000 r this reaches three and
a half days.
.The effect of still larger doses is most remarkable, since
these doses do not shorten the time for death any further.
To bring this point out we have to consider the results with
a graph plotted on quite a different scale. In Fig. 18 we were
dealing with deaths in days and hundreds of roentgen, now
we need hours and tens of thousands of roentgen. This is
seen in Fig. 19, which shows that the survival time for
irradiated mice is constant at three and a half days for doses
between 1,000 r and 15,000 r. With very high doses the
time for death falls again, and at IOO,OOO r is only one hour.
We are now approaching the region of instantaneous death;
with 200,000 r the animals die while being irradiated, even
though with exceptio’nally powerful equipment available
this amount of radiation could be given in a few minutes.
This relation between dose and survival time is not confined
to mice, and has been shown to apply to rats and guineapigs
as well, though the numerical values for the different
stages differ. The constant survival period for guinea-pigs
is four and a quarter days and for rats two and a half days.
These variations are small, and there is every reason to
believe that the response of mice is qualitatively representative
of that for mammals in general. It is obvious now
that there is no one value for a lethal dose unless one also
specifies the survival time. The period commonly chosen,
:hirty days, is very informative, since the survival time rises
sharply when doses fall below this value, and after a reducdon
by as little as one third a decrease in survival time
already becomes difficult to establish experimentally.

Different radiations.
So far only the effects of X- and y-rays
have been discussed; that is radiations giving rise to low ion
density tracks or having a low rate of loss of energy (see
p. 23). More densely ionizing radiations are more effective
in producing effects on isolated cells than radiations of low
ion density (see p. 206), and it is not surprising that this also
applies to the death of mammals. Since it is necessary to
irradiate the whole body of the animal it is not possible to
compare the effects of particulate radiations such as
a-particles or protons from an external source since these
radiations would hardly penetrate the skin because their
range in tissues is only a small fraction of an inch. Extremely
large doses of such radiation are therefore necessary to
produce radiation sickness and death within thirty days,
since only part of the body is being irradiated. These radiations
produce intense local skin reactions, but do not kill.
The same applies when animals are irradiated with Beta-rays,
which do not differ in ionization density from X-rays or
gamma-rays but which cannot penetrate deeply; with these a dose
ten times that for X-rays is necessary to kill mammals.
These findings emphasize again the great increase in resist- ,
ance to radiation which results if part of the body is unirradiated
and is able to begin repair processes.
Neutrons have a very great range, for reasons discussed
on p. 34, and can be used in the same way as hard X- and
gamma-rays to produce uniform irradiation of the whole body.
They give rise to protons inside the irradiated tissue, and
these have a much higher ionizing density than the electrons
released by the X- and y-rays used in work of this type.
No single figure can be quoted for the ratio of the effectiveness
of fast neutrons to that of therapy X-rays (e.g. 200 kV)
since different values are obtained depending on the exact
conditions of irradiation. For lethal effects at reasonable
dose rates neutrons are about three times as effective as
X-rays. In other words it is necessary to deposit only one
third of the amount of energy when this is derived from
neutrons as compared with that needed when this is derived
from X-rays.

Something approximating to a whole body irradiation
with a-rays can be achieved by causing the animal to inhale
the gas radon which emits a-particles and distributes itself
fairly uniformly through the body. In this way it was found
that this radiation was about twice as effective as neutrons
e. five to six times as effective as X-rays) in producing
acute death, illustrating the danger of densely ionizing
radiations. The body is able to recover less from the radiation
of higher specific ionization than it does after exposure
to X-rays and y-rays, and the relative effectiveness of
neutrons becomes even greater when chronic effects at low
dose rates are studied. This can be seen by comparing the
recovery in an organ with frequently dividing cells, such as
the spleen, after irradiation with X-rays and with neutrons,
with a dose which produces in each case a high proportion
of cell death. The appearance of the organ a day or so after
irradiation will be the Same, but after about a week regener-
ation and new cell colonies will be seen only in the animal
irradiated with X-rays. Consequently the destructive action
of neutrons does not fall off nearly so rapidly as that of
X-rays on decreasing the dose rate. Thus chronic effects can
be produced by repeated exposure to fast neutrons at one -tenth
or even less of the dose necessary with X- or gamma irradiation.
This applies particularly to theincidence of opacities
in the lens of the eye, eventually leading to cataract, so
safety precautions when working with fast neutrons must
be much more stringent than those for X- or y-rays. Besides
being more effective the effects of densely ionizing radiations
are sometimes qualitatively different. The symptoms
after irradiation with neutrons are not the same as those
observed after exposure to X- and y-rays, and the cause of
death would appear to be different.
Although exposure to alpha- and beta-ray-emitting isotopes is
much less hazardous than exposure to isotopes emitting
gamma-rays, because of the lack of penetration, the former
become dangerous if they are inhaled, since they may
become widely distributed and thereby give rise to a whole body
dose or become fixed in certain organs, when they
often induce cancer (see p.149). For this reason strict safety
precautions must be observed in handling radioactive
materials, even if the radiation dose received from them by
external irradiation is very small. The Beta-ray emitting radioactive
dust which contaminates enormous areas after
nuclear explosion is likely to produce more casualties than
the momentary flash of neutrons and gamma-rays released during
the explosion.


THE CAUSE OF DEATH
The symptoms of acute radiation sickness are summarized
in Table v, but these tell us little about the cause of death.
If a pathologist carries out a post-mortem examination on
a mammal which has succumbed after a radiation dose of
a few hundred roentgen he would find it very difficult to
pinpoint death to failure of a particular organ. He would
find internal haemorrhages of varying severities, and in
some cases, though by no means all, they would be considered
to be the cause of death; haemorrhage can kill either
by destroying the function of a vital organ or by producing
severe anaemia. But the site at which haemorrhage is found
varies in different animals; in rats and mice it is confined
almost entirely to the intestine, where it would not be lethal;
whereas, in the pig, haemorrhages are also found near the
heart where they could kill. Now anaemia is observed in all
animals after irradiation, though death often occurs before
this is really severe. If anaemia causes death, then blood
transfusions ought to prevent it, but in practice no increase
in survival or even significant increase in the time between
irradiation and death is found by giving extensive
transfusions.

t

After irradiation the body loses some of its ability to
produce antibodies which combat invasion by bacteria
against which it is normally immune (see p. 100) and
following an atomic explosion the surgeon finds that even
minor wounds become septic, in spite of the most rigorous
precautions. As a result, infection is almost invariably seen
after a whole body dose of a few hundred roentgen. Because
of the haemorrhages in the intestines the body is invaded by
bacteria from this source, and such infections of course are
extremely dangerous, but if they were the ultimate cause
of death, treatment with antibiotics should decrease mor-
tality. The effect of these modern drugs is dramatic wheresease
ever a disease is brought about by bacterial infection,
especially since a number of different types, capable
between them of destroying most kinds of bacteria, are now
available. Yet the effect of these drugs in radiation sickness
is from spectacular; there is some evidence that a
combination of antibiotics can produce a slight increase in
the time between the lethal dose and death, but the mortality
tality rate is not decreased; in other words treatment with
antibiotics does not increase the resistance of the animals
to a lethal dose. This proves that infection alone is not the
cause of death, but it does not mean that for people who
have been irradiated these infections need not be treated.
With marginal doses effective treatment of the infection
with antibiotics may make all the difference and it can also
change the course of the sickness; for example, after
irradiation rats suffer from severe diarrhoea, which contributes
to death because the loss of water and salts disturbs
the normal metabolism; if treated with the antibiotic,
aureomycin, the diarrhoea disappears, but the animal dies
none the less, although a few days later than it would have
done so without treatment. When newly hatched chicks are
irradiated their kidneys fail to act, and in this case it is
possible to attribute death to damage in one specific organ,
but this is very unusual. In view of these variations it may
appear that radiation sickness is an ill-defined condition
and that it should not be considered as a whole. Indeed,
some have held the view that since the symptoms and apparent
causes of death are different for different animals, the
fundamental mechanism leading to death is different, too.

No complete explanation can be given the paradox
that although the pathological changes from whole body
irradiation are diffuse and ill-defined, yet death occurs
with remarkable regularity. The problem must be different
from poisoning with chemical substances, which have to
penetrate through the body and eventually become localized
in certain organs. Initially radiation acts equally on all
cells, since it does not have to rely for its dissemination on
transport by the circulating fluids, such as the blood stream,
As a result damage is very widespread; for example, every
part of the bone-marrow is affected and no parts are available
to undertake repair; this is clearly shown by the
increase in survival which occurs when very small parts of
the body are shielded against the radiation (see p. 75),
The whole of radiation sickness is a complex interplay
between cellular damage and impaired recovery processes;
this introduces the great diversity and apparent lack of
specificity. After whole body irradiation mitosis is stopped,
blood production ceases in the bone-marrow, and the
walls of the intestine are no longer replaced by new cells as
they get worn away. Microscopic examination of these
organs will show them to be completely devoid of new cells
a day after a whole body irradiation of several hundred
roentgen. Yet this same effect is produced with doses which
do not kill, and no difference can be seen one or two days
after irradiation between animals which have received a
lethal dose and those which have been given a slightly
smaller dose which does not kill. The inhibition of cell
division is only temporary, and the bone-marrow and the
walls of the intestine begin to fill up with cells again, three
or four days after irradiation, so that at the time of death
these organs do not appear to have been severely damaged;
yet there can be no doubt that the act of putting them
temporarily out of commission contributed to death.
Presumably by the time these radiosensitive organs are
producing cells again the opportunity for repair and
recovery of other irradiated cells has been lost.
The great interdependence of the different organs and
cells in the body may be the reason why mammals are so
much more susceptible to radiations than most unicellular
organisms. The effect of irradiating one organ is often to
impair the function of another, which need not even have
been exposed to radiation. By irradiating the whole body
the balance between organs is disturbed, and a vicious
circle is set up: a disturbed condition persists for several
days, brought about largely by the fall in blood cells, which
in turn are not being replaced by the bone-marrow. By the
time the individual cells have recovered, it is too late for
delicate interplay between the different organs to be
re-established, and the complex machine comes to a standstill.
One might almost say that it is the diversity and the
absence of any specific response which is the cause of death.

Ionizing radiations hit mammals at their weakest point:
the co-ordination of function at the level where it is beyond
the control of the brain. The brain and central nervous
system are remarkably radiation-resistant, and there is no
indication that impairment of these plays any part in
radiation sickness.

EFFECTS FOLLOWING NON-LETHAL DOSES *
When considering the injuries following doses of radiation
which are not lethal, it is necessary to differentiate between
long- and short-term effects. The long-term effects are those
which become noticeable only a considerable time after
the radiation is given, and when all the immediate signs of
radiation have disappeared and are probably forgotten.
* The discussion is confined to X- and gamma-rays. Irradiation with
densely ionizing radiations produces the same symptoms but at very
much lower doses. The relative importance of the different injuries and
the time at which they appear are different for different radiations. The
densely ionizing rays such as neutrons are more effective than X- and
gamma-rays, but the magnitude of this difference – the relative biological
effectiveness (RBE) see p. 206 – is not the same for different effects produced.
As already mentioned (see p. 81) the RBE of densely ionizing
particles is greater when the comparison is made at low as opposed to
high dose rates.

These delayed effects may also arise after periods of continued
exposure to radiation at a low intensity, so that no
immediate symptoms become apparent. Besides the time
factor there is another characteristic difference; the damage
giving rise to immediate effects – radiation sickness – is
quickly repaired by the body and the severity of the symptoms
is therefore very dependent on the rate at which the
radiation is received. The injuries which become apparent
as late changes seem to be irreparable, or at best only
incompletely reparable, and are therefore much less dependent
on dose rate. This is the reason why radiation at very
low intensity may fail to produce any symptoms typical of
radiation sickness, while still bringing about long-term
injuries, some of which will be dealt with in the next two
chapters.
Radiation sickness can be of two types: the one follows a
single large dose and the other small doses repeated over
long periods. The so-called critical dose in a single irradiation
is about 100 r, and below this no unpleasant symptoms,
such as vomiting and general lethargy, are observed. Blood
tests (see below) can reveal a single exposure between 25
and 50 r, but below this dose no pathological changes can
be found. Symptoms of radiation sickness also occur after
repeated radiation with low intensities, although the appearance
of long-delayed effects complicates the picture. From a
study in Britain, Sweden, and the U.S.A. of hospital personnel
connected with radiation therapy, it would appear
that a dose of less than I r per week produces definite,
though slight, symptoms after several years. However, the
effect is small and in animals any harmful effects can only
be recognized following continual irradiations with more
than 5 r per week. If this dose is given throughout life a shortening in
life-span in mice can just be detected, although much
higher doses are necessary for this to become marked (see
Fig. 17).

The safety or tolerance doses are not dictated by
considerations of radiation sickness, but rather by the
dangers from long-term effects (see Chap. 4). This is the
reason why the internationally accepted level of X- and
gamma-radiation to which research and industrial workers and
hospital staff can be exposed without harm is set as low as
0.3 r in any one week, and it has been suggested that the
tolerance dose should be reduced to 0.1 r/week. If a
particularly hazardous operation results in the worker
getting, for example, 0.2 r/hour, then he can only do
this work for fifteen hours in the week and for the rest of
the time must stay away from all radiations. In Britain
only a very small proportion of those engaged on work
involving substances or machines which give off atomic
radiations receive the full tolerance dose, and the majority
of the staff at Harwell, for example, receive less than one
tenth of this dose.

For densely ionizing radiations, such as neutrons or
X-rays, no definite safe level has been proposed, since the
data on which a recommendation could be based have not
yet been obtained. We do know that they are much more
effective in producing many types of injury than the sparsely
ionizing radiations, and at present one prefers to err on the
side of safety and set a tolerance equivalent to 0.01 r per
week for these radiations.

Course of illness. When a single heavy dose is received, as it
was at Hiroshima and Nagasaki, the victims can be divided
into three groups,* and the course of the sickness for these
* The same sequence of events was also found with the Japanese
fishermen who accidentally received lethal doses from the radioactive
fall-out following the test explosion of the first hydrogen bomb, the force
of which greatly exceeded expectations. Very much has been learned
about the details of severe radiation sickness from the clinical history of
eight physicists who received large doses of radiation in an accident
which occurred at the American atomic research laboratories at Los
Alomos. In an experiment to demonstrate that an explosive chain reaction
sets in when two pieces of polonium exceeding a critical size are
brought together, the hands of Dr Louis Slotin slipped, and he and his
colleague received whole body doses exceeding 600 r which ended in
death on the ninth and twenty-fourth day. Bystanders who received a
smaller dose recovered. In spite of the most intense efforts made to save
them nothing could be done to stave off the total collapse of the two
scientists who died. The progress of the disease could not be influenced
by any treatment.

categories is summarized in Table v, which applies when
there are no complicating factors due to burns or other
injuries. The most remarkable feature of the whole illness
is that, apart from attacks of severe vomiting which set in
within an hour or so after irradiation, the victim has no
indication that he has received a fatal or near-fatal dose.
If a dose in excess of 600 r has been received death is almost
certain, and a few days after the dose the patient is obviously
very ill. His temperature rises and he loses weight very
rapidly because he can take no nourishment. The linings of
the stomach have been so severely damaged that food can
no longer be absorbed. Within two weeks or a month at
most death occurs, and so far there is no treatment which
can be given to human beings which alters the progress of
these events. The methods of assisting recovery, described in
Chapter 7, are not sufficiently advanced to be used clinically.
If the vomiting stops after a day or two the dose received
is in the lethal range; that is, a considerable proportion, but
not all, of those affected will die (dose about 300 to 500 r),
but the outcome is not necessarily fatal. For one or two
weeks there are no well-defined or distressing symptoms,
and even the general feeling of lassitude which sets in a day
or so after the irradiation may wear off. But quite suddenly
a variety of signs due to infection and anaemia are observed,
as well as loss of hair and-diarrhoea. From then on the
patient may go steadily downhill or may recover, although
he will not feel completely fit for some months. Prompt
treatment of infections probably increases the chance of
survival, but on the whole the outcome is almost entirely
dependent on the make-up of the individual, and apart from
rather obvious medical measures there is little scope for the
physician to influence the course of the illness.

Doses of the order of 200 r produce no immediate symptoms,
and if uncomplicated by other factors should not
prove fatal. After two or three weeks a general feeling of
ill-heath becomes apparent, followed by other symptoms of
a more definite type. Anaemia will take longest to disappear,
and it may be several months before complete recovery

has been achieved; 100 r or less does not bring about any
obvious symptoms of illness and only a blood count shows
that the irradiation has occurred. A single dose of the order
of 25 r produces no observable effects, and can probably be
accepted as a calculated risk, under certain conditions.
Fortunately no data are available for man of continued
relatively high dose levels, and human
experience of protracted exposure is limited to doses which
give rise at the most to changes in the blood picture, but
none that cause death by radiation. Experience with experimental
animals indicates that with continuous irradiation with more than 50 r
per day the symptoms are the same as for a single large dose, except that
they will set in after many weeks when the total dose received is several times
the lethal dose for a single irradiation.

Smaller daily doses right down to 2 or 3 r per day can still be considered
lethal, since they definitely shorten life, but the illness now has a quite
different character. None of the typical symptoms such as
diarrhoea, severe anaemia, or loss of appetite and weight,
are observed, and the changes seen can best be summarized
as being akin to those normally associated with old age.

Continuous irradiation with doses which do not exceed 25 r
per day may be said to age the animals prematurely. The
data in Fig. 17 can be taken to show that the lethal
dose (LD50) of X-rays given at the rate of 10 r per day is
6,500 r. A more useful way of looking at the results is that
continued irradiation of 10 r/day lowers the life-span
(expressed as the period in which half the animals have
died) of the rat from ninety-five weeks to sixty-five weeks.

In the same series of experiments I r per day lowered the
time necessary for half the animals to die to seventy-eight
weeks. Even as little as 0.1 r per day may have produced
some reduction in life-span.

A large single dose which does not kill also leaves a permanent
mark which is revealed as premature ageing. There
is an increase in the incidence of cancer a long time after
irradiation (see Chapter 5) and this will contribute to a
decrease in life expectancy. But even when allowance
has been made for such deaths, irradiated animals
still die sooner than normal ones (see Fig. 20). No
specific cause for these earlier deaths can be given and they
follow the same pattern as those of ordinary animals dying
f
of old age. Some investigators believe that atomic radiations
hasten the onset of typical senile alteration and can be
considered to accelerate ageing.
Efect on the blood. There are two circulating fluids in
mammals which are responsible for supplying every part
of the body with food derived from the digestive organs and
with oxygen from the lungs. In addition they have to dispose
of waste products, which are eventually eliminated by the
excretory organs; and of carbon dioxide which is exhaled.

The blood circulates through a closed system of tubes and
is pumped very efficiently by the heart. The whole round
trip through the body and lung and back to its startingpoint
only takes about two minutes. All tissue is bathed in
fluid which exudes from the finest capillaries carrying the
blood and is collected again and returned to the veins by
lymph vessels which ensure the forward flow of the liquid.
To fulfil its purpose the blood contains cells which originate
in the blood-producing organs from which they are discharged
into the blood stream where they perform different
specific functions. After a certain time the blood cells
deteriorate and are then disposed of by the spleen or alimentary
canal. Under normal circumstances the blood cells are
replaced as required by the blood-forming organs which
carry, however, a reserve supply for release in emergency .
such as a haemorrhage. The volume of tissue fluid is kept in
balance by the intake of drink and its distribution between
the blood and tissue is complex, involving a number of
factors and organs.
The red cells are by far the most common, and there are
about 500 of these t; every one of the others, which are
known as white cells and consist of many different types.
The main division is into lymphocytes, which are made in
the lymph glands, and granulocytes, which are produced in
the bone-marrow: the latter can be divided into a number
of sub-groups, the most common of which are the neutrophils.
Since lymph tissue is spread throughout the body,
nothing short of total body irradiation will influence all the
sites at which they are produced, but for the same reason it
is almost impossible to irradiate any part of the body without
exposing some lymphoid tissue. Lymphocyte counts
are therefore a very sensitive biological indicator for detecting
irradiation. The red cells and the various types of
granulocytes are formed in the bone-marrow and a few other
organs, from one and the same basic cell, the reticulocytes.
This is the precursor of a number of cell types and environmental
conditions determine the kind of differentiation they
undergo to become a particular type of mature blood cell.

Once they have acquired their full degree o f specializa~io~\
they are ko longer cipable of division,&d new cells neecletl
for replacement are produced by division of the retic~tlocytes.
The lifetime of the cells varies greatly; lymphocytes SUP
vive for less than a day; granulocytes live for about thrw
days; while the same red cells carry on for three or four
months before they break up and have to be replaced. A
short-lived blood component of the greatest importance is
the platelets, which are related to the granulocytes and
carry enzymes that are necessary if blood clotting is to occur.
Platelets are therefore vital for the control of haemorrhages
and if there are too few present, small injuries – particularly
internal ones – which would normally be quite harmless,
may prove fatal.
Radiation to the whole body affects a11 the different cells
of the blood, but as with all radiation injuries these changes
f
only become noticeable some time after irradiation. The
general pattern for a dose which does not kill is shown in
Fig. 21. A reduction in the number of lymphocytes can be
recognized less than one hour after irradiation. In addition
to giving a very early indication, a significant drop in
lymphocytes is found in man and animals after as little as
one dose of 25 r of total body irradiation. This great sensitivity
is due to the fact that the circulating lymphocytes have
such a very short life, and any intederence with cell division
in the lymph-forming tissue is bound to be reflected very
quickly in the blood picture. In addition the lymphocytes
themselves are very radio-sensitive and radiation may
shorten their life and thereby aggravate the shortage resulting
from the hold-up in mitosis. With less than lethal doses
a gradual upward trend in the lymphocyte count can be
observed after three days, but the subsequent rate of recovery
is rather slow. For doses in the lethal range where some
of the animals survive, the lymphocyte count does not begin
to recover for about two weeks and in the early stages gives
no indication whether an animal will survive or not. The
low number of lymphocytes is not thought to be critical to
the animal and does not contribute to the more serious
aspect of radiation sickness, but it is diagnostically most
useful.
The fall in granulocytes can be observed only a day after
the irradiation and the lowest value is reached after about
seven days. The dose required to give a significant depression
lies between 50 and 75 r and it is therefore less radiosensitive
than a lymphocyte count, but it provides a much
more important indication of the seriousness of the radiation
injury. The number of granulocytes does not return to
normal for at least two to three weeks, and it is during
this period that death from acute fadiation sickness occurs.
An increase in the number of blood cells is a very good
sign, and suggests that recovery will occur. With chronic
exposure the fall in granulocytes occurs very slowly, but
there is also much less power of recovery than with Iymphocytes,
and the count may stay permanently low after all
exposure to radiation had come to an end.

The restoration of the granulocyte count following single
irradiation is brought about by the release of cells held in
reserve in the spleen, and in addition reticulocytes are
released from the bone-marrow before these have had a
chance to develop into the fully differentiated cells. Tlir
presence of the so-called immature cells is very typical 01′
radiation damage and about two to four weeks after a sevcrr
exposure (depending on the animal used) the number ol’
these cells is greatly in excess of those normally present
(Fig. 22). The presence of these immature cells is not neces-
f
sarily beneficial, since they cannot exercise all the functions
of the fully differentiated cells and can be harmful to some
important organs. This is a typical case of how radiation
can upset the fine balance of the whole body and interfere
with organs which are themselves radiation-resistant.
The fall in platelets can be observed at about the IOO r
level and in general follows that of the granulocytes, except
that recovery occurs rather more slowly. The fall in platelets
is of the greatest importance, since it is responsible for the
haemorrhages which are such a characteristic feature of
radiation sickness. Increased fragility of the blood vessels
after irradiation is a contributory factor, but the most
important cause of haemorrhage after irradiation is the
reduction of platelets and the consequent failure of the blood
to coagulate. The severity and most common site of these
haemorrhages vary from species to species and depend
on the dose. Although they are unlikely to be the cause of
death following an irradiation with a dose sufficient to kill
all the animals (i.e. substantially greater than LD~Ot)h,e
severity of the haemorrhages may often determine whether
an animal lives or dies after a dose near the lethal range.
A little extra exertion or a small additional injury, such as
a cut or abrasion, may greatly affect the severity of the
haemorrhage because of the virtual absence of platelets.
The red cells themselves are extremely radiation-resistant,
and doses of tens of thousands of roentgen are necessary
before any change can be detected in their behaviour.
Exposures of the order of those which produce radiation
sickness will leave the circulating cells entirely unaffected
and allow them to continue their normal function. Since
they persist for many months, interference with the formation
of their precursors by stopping cell division in the bonemarrow
cannot make itself felt for many weeks, by which
time normal cell division – as shown by the recovery in the
lymphocyte count – will have set in. Nevertheless a diminution
in the red blood cells is observed after whole body doses,
about one week after irradiation, and continues for about
three weeks; it is at a minimum when the lymphocytes and
granulocytes are nearly back to normal. The reason for this
drop and the general anaemia always found after severe
irradiation is the haemorrhages, which often result in
serious loss of blood. Chronic irradiation may give rise to
protracted anaemia, even though the white cells are normal,
because the areas of the bone-marrow in which the red cells
are formed have less power of recovery from radiation
damage than those producing white cells. With a single dose
permanent loss of generative power is generally small and
the immediate effects on the blood are due to temporary
inhibition of cell division, resulting in a deficiency, followed
by release of immature and abnormal cells, which tend to
upset some of the many functions for which the blood is
responsible. The status quo is eventually re-established if the
dose has not been too great. Protracted irradiation may
effect an irreversible change by permanently reducing the
number of blood cells produced.
f

to lose weight (see Fig. 23) because of the general symptoms
of severe radiation sickness which lead to a complete loss of
appetite and to diarrhoea. Animals which do not survive
continue to lose weight until death, but the curve for
survivors passes through a minimum, at a time which
depends on the species and the severity of the radiation, and
then goes up again to that before irradiation. If young
animals which are still growing are used, a second weight
drop can be observed. This is clearly brought out in Fig. 24,

f

where it can be seen that the initial weight loss follows the
lymphocyte count and is reversed on the general recovery
of the animals within the first week. Then suddenly, about
two weeks later, a second drop occurs which coincides with
the fall in the number of red cells, and this weight loss is a
measure of the haemorrhages. If these haemorrhages nrp
very severe because of some additional factors, death will
occur at this time. That is, the animal has survived the first.
shock of the radiation which normally kills at lethal doses,
but its platelet shortage has prevented it from surviving the
subsequent haemorrhages.
This explains why there are two distinct periods of death
after irradiation with a dose which is in the lethal range but
not invariably fatal. There is death after about ten days
following the initial weight loss – this is the period in which
death occurs with doses sufficient to kill all animals – and an
occasional death after two to three weeks. In man very little
can be done to prevent death in the first period (see p. 84) ;
but these late deaths can be prevented by medical treatment
for haemorrhages, since the number of platelets will be going
up again, parallel with the recovery of the white cells in
general.
Infection. Healthy mammals have the capacity for forming
substances, known as antibodies, which combine with and
thereby put out of action foreign bodies which enter the
system. This is one of the major defences of mammals
against disease-producing organisms. When, for example,
bacteria from food which has gone bad are absorbed, antibodies
against them are made by the animal and will keep
the infection in check so long as the invaders are not too
virulent and do not damage the host too severely before
sufficient antibodies have been made. Immunization consists
of stimulating the body to make antibodies against a
particular bacterium or virus which is dangerous. This is
usually done by injecting the virus in an attenuated form, in
which it still stimulates antibody formation but does not
produce the symptoms of the disease (compare p. I 99). After
immunization the body has the antibodies in readiness, or
the antibody-forming organs have ‘learned’ how to make the
particular antibody required and can produce it more
rapidly the next time. Then when it comes into contact with
the invader again the immunized animal is able to combat
lie dangerous organisms immediately, before they have a
rhance of becoming established.
One of the most typical symptoms of radiation sickness is
the much greater susceptibility to infection. Small and
insignificant wounds turn septic after a whole body irradiation
and many infections set in spontaneously. The reason
for this is twofold; the mucous membranes of the intestine
become thin and ulcerated and aided by local haemorrhages
the bacteria from the intestine gain access to other parts of
the body. This in itself would probably not be sufficient to
make the infections as serious as they are, but the body has
at the same time lost its capacity to form antibodies. These
are probably produced inthe blood-forming organs, and
interference with them is another aspect of the effect of
radiation on the blood. Whether the cells which actually
make the antibodies are destroyed and become replaced by
new ones, or whether radiation merely temporarily inhibits
their protein-synthesizing activity, is not known. However,
the dose required before loss in immunity-producing pro-
~ertieiss observed is of the order of several hundred roentgen.
Shielding part of the blood-forming organs from the
radiation prevents the interference with antibody formation,
and this must contribute to the great protective effect of
excluding parts ofthe body from irradiation (seep. 75). These
infections are a very serious part of radiation sickness, and
if not treated can lead to death, but they are not the normal
cause of death with doses of radiation in the lethal range.
Effect on reproduction. It is necessary to distinguish between
localized irradiation confined to the sex organs, and whole
body irradiation. Permanent sterility can be produced by
irradiating the ovaries with about I,OOO r; in the male this
dose will give only temporary sterility, and permanent
destruction of the organs in which the sperms are generated
requires several thousand roentgen in most mammals,
including man.
Following a whole body irradiation of 400 r or more the
male remains fertile for many weeks, but then becomes
sterile for a protracted period of several months. Recovery
of the reproductive organs takes longer than that of othn’u,
and sterility can be said to be the most protracted symptom
of radiation sickness. The reason for the delay of about two
weeks between irradiation and the diminution in tlir
number of sperms – complete sterility requires a minimum
of 400 r and occurs as late as a month after a single dose of
radiation – is due to the fact that the mature sperms arc
very radiation-resistant, so that successful mating is possible
by a male who has received a lethal dose and will shortly
die. The earlier stages of the germ cells are radiation-sensitive,
so that the supply of mature sperms is interrupted by
the irradiation. The long delay before fertility occurs again
is due to slow recovery of the germ cells coupled with the
fact that it takes several weeks for a newly formed germ cell
to develop into a mature sperm. Complete recovery occurs
in animals and, so far as can be judged from the limited
number of ‘atomic accidents’ and the Japanese casualties,
in man also.
Ovarian tissue is much more sensitive and it has been
claimed that a whole body irradiation of 150 r has caused
permanent sterility in mice. Again the generative cells are
the most sensitive, while the fully differentiated ovum is
unaffected. Menstruation will therefore continue for a time
even after a very heavy dose (with mice two litters can
occasionally be got after irradiation before permanent
sterility becomes apparent), but will cease for a long period
or completely if the dose exceeds the so-called castration
level (about 300 r for human beings),* which is however
rather variable even within one species.
The skin. In March 1896, only three months after Rontgen
published his discovery of X-rays, their destructive action
* The practice of gynaecologists of stimulating fertility by irradiating
the ovaries with 175 r is no contradiction of this, since this is a localized
exposure as opposed to a whole body dose. Reference is made on p. 125
to the harm which could occur if this form of treatment were carried out
widely.

on living matter was revealed by their depilatory action.
Irradiation with a few hundred roentgen by all types of
rays causes hair to fall out about a week or so later. This
is a local effect and the whole body need not be exposed:
the part irradiated can be recognized from the loss of
hair, This is particularly noticeable following exposure
to &rays, given off by a radioactive substance, since these
only penetrate a small fraction of an inch and do not, unless
ingested, produce general radiation sickness. Usually the
hair growth returns and no permanent damage is left
behind with doses of two or three hundred roentgen. This is
illustrated by the photographs on Plate ga, which show a
young South Sea Island girl, on whose head had settled
radioactive dust released in an American hydrogen bomb
trial. A year later the hair had returned to normal. For most
human beings more than 500 r permanently damages hair
growth and the new hair will be sparse and weaker. Larger
irradiations still cause permanent baldness; but for all these
changes it is difficult to give a definite threshold dose, since
there are remarkable variations not only between species
but even from one person to another.
After fairly big doses the appearance of the hair and
sometimes also its colour are altered. These changes are
permanent, indicating that cells responsible for secreting
the pigment or other substances associated with the hair
have been permanently destroyed. An effect of this type is
shown most markedly in certain strains of black mice,
whose hair is turned permanently white after irradiation;
all new hairs in the irradiated region are white, and the
pigment never returns. Plate 8a shows the greying produced
by irradiating a mouse with 800 r of X-rays in a narrow
beam. The X-rays are very penetrating and go right through
the mouse, so that the white colour occurs on both sides of
the animal. Since this dose was confined to a very small
volume only, no symptoms of radiation sickness were
produced.
The effect of radiation on the skin is extremely well
defined, and was used as a means of measuring the dose
from X ray machines before physical methods had been
developed. The changes produced depend only the local
dose received, and are independent of radiation to the
remainder of the body. Damage to the cells in the skin is not
influenced by irradiation of other organs or of other skin a1
some distance away, and differs very markedly in this
respect from irradiation of the blood-forming organs and
the associated general symptoms of radiation sickness.
Injury follows a general pattern; within a few hours after
exposure the affected parts redden temporarily. This effect
is usually very mild and sometimes is not observed at all;
it wears off after a day or so. More than a week later definite
reddening develops and the extent of this so-called erythema
varies very much with dose and dose rate. This reddening
is an inflammation of the kind which is evoked in damaged
tissue by many forms of injury, not only radiation. The skin
shows remarkable power of recovery, so that by extending
the time over which the dose is given to more than one hour
the damage is greatly decreased; as the dose rate decreases,
the total dose necessary to produce a certain degree of
erythema is increased. This is shown in Table VI, which
gives the total dose necessary to produce in man a fixed
intensity of skin reddening at two weeks following the
irradiations. Narrow beams were used, so that the total
areas involved were very small.
t
The dose at which definite erythema sets in is usually the
same as that causing loss of hair; it is quite temporary and
causes little distress. It is often followed by slight changes
in pigmentation, and large freckles may be seen. In coloured
races loss of pigmentation may occur (see Plate 9b).

With doses exceeding I,OOO r the erythema becomes
severe, the skin becomes dark red and blisters are formed.
‘I’hese indicate extensive damage to the skin and can turn
into extremely unpleasant running sores which take months
to heal and leave bad scars. In radiation treatment of deep-
seated cancers the dose tolerated by skin without producing
severe erythema is often the limiting factor. Accidental
irradiation and contact with radioisotopes have given rise
to severe skin damage, which clears up quite quickly if the
exposure is not repeated. Plate 9b shows the feet of a native
from the Marshall Islands, which were contaminated by
radioactive fall-out. Some months later they were completely
healed. (Paul’s note: this is an overly optimistic assessment, as later
history reveals, and indeed as the author notes in the captions to
the photographs.)

The effect on the skin of long-term irradiation, such as
exposure of their hands by the early radiologists, is much
more serious. A complex process of damage followed by
periods of repair often results in overgrowth and causes the
skin to dry up. Gradually and insidiously the fingers become
stiff. Once this has occurred recovery becomes impossible,
and even if all further exposure to radiation is then avoided
the affected parts will not improve and may continue to
deteriorate. Typical deformities following prolonged local
exposure to large doses, which did not produce radiation
sickness since only a part of the body was exposed, are
shown in Plate 8b.

The first warning that irreversible damage is being done
to the skin by irradiation over long periods is a change in
the ridges of the finger-tips. This test can be made very
sensitive by taking finger-prints which immediately reveal
the flattening or disappearance of ridges (see Plate 10a). As
the skin is damaged it is replaced, and even when the skin
peels off the fingers the identical finger-print pattern is
re-established. With chronic radiation a time arrives when
the repair process fails to produce an exact copy and the
new pattern remains permanently. Even today many radiotherapists
and surgeons, who take reasonable precautions, reveal their profession
by their finger-tips; certainly the surgeon who inserts radium needles cannot
fail to expose his fingers and must watch the symptoms most carefully,
for radium damage may affect the agility of his fingers on
which his skill depends.

Cataract. One of the most serious of the non-fatal consequences-
of irradiation is the clouding over of the lens of the
eye. This is known as cataract formation, and can lead to
total blindness, although frequently vision is only impaired.
There is a long latent period, and five to ten years may
elapse between irradiation and the appearance of symptoms.
The most important factor is the irradiation of the eye
itself, and exposure of the rest of the body is not important.
In many respects the irradiation effects of the eye follow
closely those of the skin. Dose rate is more important than
total dose, and a single heavy dose is the most effective for
inducing cataracts. About 400 to 500 r of X- or y-rays are
needed, and for all practical purposes it is not therefore a
hazard. Some of the seriously affected victims of the atom
bomb explosions in Japan who survived a big dose later
developed cataracts; in general, the threshold dose for
serious cataract is close to the acute lethal (1-050) dose for
whole body irradiation.
Radiations giving rise to densely ionizing tracks, such as
fast neutrons, are generally more effective in inducing all
types of radiation injuries than X- and y-rays (see p. 206),
but they are exceptionally more effective in producing
cataract. Thus the neutron dose necessary for cataract lies
well below the LD50 for whole body irradiation. This was
first indicated in 1948, when several cases were found in
physicists who had been exposed to neutrons from a cyclotron.

The greatest danger comes probably from accidental
exposure to Beta-ray-emitting isotopes, whose radiations do not
penetrate and do not therefore produce whole body effects;
fall-out from bombs is an obvious source.

BREAKING NEWS! (US) HOUSE PASSES CONTINUING RESOLUTION WITH $7 BILLION IN TAXPAYER LOANS FOR NEW REACTOR CONSTRUCTION

December 10, 2010

BREAKING NEWS!

HOUSE PASSES CONTINUING RESOLUTION WITH $7 BILLION IN TAXPAYER LOANS FOR NEW REACTOR CONSTRUCTION

URGENT: TELL SENATE TO REJECT ANY NEW TAXPAYER LOANS FOR NEW NUKES

December 9, 2010

It’s my birthday, and I wasn’t planning on having to send an Alert out today, but…

The House last night approved a Continuing Resolution to fund the government through next September that adds $7 Billion in new taxpayer loans for new reactor construction. They also threw in $3 Billion for “clean” coal.

That’s $2 Billion less than the most recent Department of Energy request, and is less than 20% of what President Obama asked for in February–but it’s still $7 Billion too much!

The Senate will vote on this as early as tomorrow. Please send a letter to your Senators today here. (US eligible people only) If you can, please also call them and tell them no taxpayer dollars for nuclear reactors! Their direct phone numbers will appear when you send your letter.

And please, after you send your letter you’ll be directed to our donation page. If you haven’t contributed yet this holiday season, please make a tax-deductible donation of any size–it really makes a difference and makes our work possible.

Thanks for all you do!

Michael Mariotte
Executive Director
Nuclear Information and Resource Service
nirsnet@nirs.org

Paul’s note:

SUPPORT PRIVATE ENTERPRISE: SAY NO TO NUKES.
NUCLEAR INDUSTRY: THE WORLD’S LONGEST RUNNING ‘EMERGENCY FUNDING PROGRAM’ 1942-2010.